“The biogeochemical cycle of mercury
in the Augusta Bay”
PhD Student
Maria Bonsignore
DIPARTIMENTO DI SCIENZE DELLA
TERRA E DEL MARE “DiSTeM”
Dottorato di Ricerca in Geochimica
Ciclo XXIV
2010/2011
Tutor
Prof. Paolo Censi
University of Palermo
Co-Tutor
Dott. Mario Sprovieri
IAMC- CNR Capo Granitola
Coordinator
Prof. Francesco Parello
University of Palermo
In collaboration with the Institute of Coastal Marine Environment, IAMC-CNR of Capo Granitola
Project: “Dinamica dei processi di evasione, trasporto e deposizione del mercurio nell’area
industrializzata della Rada di Augusta e definizione delle mappe di rischio sanitario per le
popolazioni residenti”
Reviewers
Dott. Stefano Covelli,
Dipartimento di Matematica & Geoscienze (DMG) Università degli Studi di Trieste
Prof. Pierpaolo Zuddas
Université Pierre et Marie Curie, Paris-Sorbonne Institut des Sciences de la Terre de Paris.
A Mattia
Una buona regola di vita è avere sempre il cuore
un po' più tenero della testa.
Alexander Graham Bell
Table of content
Abstract.................................................................................................................................1
CHAPTER I: INTRODUCTION
Preface..............................................................................................................................7
1.1 Goals
1.1.1 General goal..........................................................................................7
1.1.2 Specific goals.........................................................................................8
1.2 Mercury element................................................................................................9
1.3 Chemical properties..........................................................................................9
1.4 Sources and application..................................................................................10
1.5 The biogeochemical cycle of mercury in the environment....................12
1.5.1 How does mercury enter the food chain?..............................................15
1.6 Health concerns................................................................................................16
1.7 Mercury in Mediterranean Sea.....................................................................17
1.7.1 The Mediterranean geochemical anomaly...........................................17
1.7.2 The Mediterranean global circulation.................................................19
CHAPTER II: THE STUDY AREA
2.1 The Augusta bay................................................................................................22
2.1.1 Geographical features...........................................................................22
2.1.2 Geological feature.................................................................................23
2.1.3 The Augusta harbor……........................................................................24
2.1.4 Industrialization history………………..................................................25
2.1.5The chlor-alkali plant in Augusta Bay……………………………........27
2.2 Human health and environmental concerns in Augusta Bay...............29
2.3 Environmental Hg pollution in Augusta Bay...........................................29
CHAPTER III: METHODOLOGY
3.1 Sampling strategy..............................................................................................33
3.1.1 Sediment sampling................................................................................35
3.1. 2 Seawater collection...............................................................................36
3.1. 3 Fish sampling.......................................................................................38
3.1.4 Atmospheric GEM assessment..............................................................40
3.2 Flux at the interfaces........................................................................................41
3.2.1 Fluxes at the sediment-seawater interface...........................................41
3.2.2 Fluxes at the air-seawater interface.....................................................44
Table of content
3.3 Samples preparation and analysis..............................................................45
3.3.1 Sediment treatment for THg analyses...................................................45
3.3.2 Sequential extraction procedure ..........................................................47
3.3.3 Fishes treatment for THg analyses.......................................................48
3.3.4 Seawater treatment for THg and DHg analyses...................................50
3.3.5 THg analyses and data quality............................................................50
3.3.6 REEs treatment and analyses in seawater............................................53
3.3.7 Hg isotopes analyses.............................................................................54
CHAPTER IV: RESULTS
4.1 THg distribution in sediment and grain size composition...................59
4.1.1 Sequential extraction procedure...........................................................64
4.2 THg content in fishes.....................................................................................65
4.2.1 Total mercury concentrations in fishes ................................................65
4.2.2 Biological features.................................................................................72
4.3 Mercury (THg and DHg) in seawater.......................................................74
4.4 Atmospheric GEM distribution.................................................................78
4.5 Hg fluxes at the interfaces............................................................................79
4.5.1 Meteorological pattern of the area....................................................79
4.5.2 In situ benthic fluxes.............................................................................80
4.5.3 Air-sea GEM Flux...............................................................................83
4.6 REEs distribution in seawater ....................................................................84
4.7 Hg isotopes analyses ......................................................................................87
4.7.1 Hg isotopic composition of fishes.........................................................87
4.7.2 Sediment Hg isotopic fractioning.........................................................89
4.7.3 Hair Hg isotopic composition..............................................................91
CHAPTER V: DISCUSSIONS
5.1 Pattern of Hg distribution in the Augusta Bay.........................................94
5.1.1 Spatial Hg distribution of THg content in sediment..............................94
5.1.2 Spatial Hg distribution of THg content in seawater..............................95
5.1.3 Mercury bioaccumulation pattern in fishes..........................................96
5.2 Toxicological effect on fish compartment...................................................98
5.3 Sediments as primary source of Hg............................................................100
5.4 Mass-balance.............................................................................................. ......102
5.5 REEs distribution in seawater....................................................................107
5.6 Application of Stable Mercury Isotopes to Biogeochemistry..............109
5.6.1 Hg isotopes fractioning in fishes........................................................110
5.6.2 Hg isotopes fractioning in sediments...................................................111
5.6.3 Hair Hg isotopic composition.............................................................112
5.6.4 Sources of Hg for fishes.......................................................................112
Table of content
5.6.5 Sources of Hg for human....................................................................113
CHAPTER VI:
CONCLUSIONS.................................................................................................116
References ..............................................................................................................118
Acknowledgement.................................................................................................132
Attachment I:
Bonsignore M. , Salvagio Manta D., Oliveri E., Sprovieri M., Basilone G., Bonanno A.,
Falco F.,Traina A., Mazzola S. (2013). Mercury in fishes from Augusta Bay
(southern Italy): Risk assessment and health implication. Food and Chemical
Toxicology 56 (2013) 184–194
Attachment II:
Bagnato E., Sproveri M., Barra M., Bitetto M., Bonsignore M., Calabrese S., Di Stefano
V., Oliveri E., Parello F., Mazzola S. (2013). The sea–air exchange of mercury (Hg)
in the marine boundary layer of the Augusta basin (southern Italy): Concentrations
and evasion flux. Chemosphere 93, 2024–2032.
Table of content
1
Abstract
Mercurial, the metaphor for volatile unpredictable behavior, aptly reflects
the complexities of one of the most insidiously interesting and scientifically
challenging biogeochemical cycles at the Earth’s surface. At the base of this
toxic metal cycle there is the conversion between the different Hg chemical
species, in which the balance between the reduced and oxidized forms
depends primary on redox system conditions.
The potential risks of human exposure to Hg, especially in the form of
monomethylmercury (MMHg), particularly prenatally, and the potential
deleterious ecological consequences from localized to global scale Hg
pollution, have given much impetus to mercury studies and regulatory
activities at international level. Much of this advancement has come since
the early 1970s, and the growth in mercury research continues at breakneck
pace. The menace of this item for environment and human health deserves
further information concerning the geochemistry of mercury, especially in
coastal marine system.
The Augusta Bay is a semi-enclosed marine area, located in the SE of Sicily
(southern Italy), well-known because of the high Hg pollution. The area
indeed has experienced, since the early 60s, a significant industrialization
phase that put in several chemical and petrochemical plants and oil
refineries resulting in a severe pollution of the surrounding environment. In
particular, the petrochemical district of Augusta Bay hosted one of the most
important chlor-alkali plant in Italy, that produced chlorine and caustic soda
by electrolysis of sodium chloride aqueous solution in electrolytic cells with
a graphite anode and metallic mercury cathode. Uncontrolled chemical
discharge of Hg occurred in the Augusta Bay until 1978, when restrictions
were imposed by the Italian legislation. For this reason, in the last decade,
several studies have provided detailed information on the pollution levels
and risks for human health of resident populations of Augusta Bay. The
effects of this indiscriminate Hg discharge include the alarming high
concentrations of the
element recently measured in sediments of the basin, prompting the Italian
government to include the Augusta basin in the National Remediation Plan.
The “Augusta case“ menaces to not remain confined to a “local problem”,
but to became a large-scale threat.
Abstract
2
Indeed, the effects of meso-scale circulation of the Ionian Sea create a
higher potential risk for HgT contamination of the basin, being affected by
the transit and transformation of the major water masses, which regulate the
general thermohaline circulation in the upper, intermediate and deep layers,
respectively. Owing to the geographical location of the Augusta basin, its
outflowing shelf waters are immediately intercepted by the surface Atlantic
Ionian Stream (AIS) and mixed with the main gyres of the eastern
Mediterranean Sea, thus representing a risk for the large-scale marine
system. This complex water circulation system, together with the closeness
with the steep continental slope (part of the Malta escarpment), make the
area a potential point sources of mercury for the entire Mediterranean sea, as
previously speculated by Sprovieri et al. (2011).
All this features make the Augusta Bay an ideal natural laboratory for
deeper insights on the biogeochemical cycle of mercury in a coastal marine
environment and the need to investigate the large-scale effect of Augusta
Bay pollution has become imperative!
With the aim to fill this requirement, an integrated model on the
biogeochemical cycle of Hg has been created.
Hg cycle is a very articulated topic. Once introduced in the aquatic system
environment the fate of Hg in the marine system is affected by sorption/
desorption processes onto suspended particulate matter and, based on
associated kinetics, it may be partially transferred from surface waters to
bottom sediments. Microorganisms, at the water/sediments interface such as
sulfate reducing bacteria (SRB), mediate the transformation of inorganic Hg
to MMHg with high rates of methylation favored by the presence of high
content of organic matter under reducing environmental conditions.
Therefore, sediments are considered key contributors of MMHg to the
marine ecosystem. Clearly, this analysis stresses the necessity for better
knowledge of the specificity of the mercury biogeochemical cycle in this
particular environment through the gathering of more data on the
distribution and fluxes among the various compartments including the water
column, sediment, atmosphere and biota.
This multidisciplinary approach offers a nice opportunity to explore the
biogeochemical dynamic of mercury in highly complex coastal areas under
important anthropic impact and the potential on larger scale diffusion.
Multiple oceanographic cruises, realized during 2011-2012 period,
permitted to collect samples of sediments, seawaters and fishes inside and
outside the Augusta Bay.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
3
Furthermore in order to trace the entire chain, from sources (polluted
sediments) to sink (man), analysis of Hg in fishes (the main route of Hg
uptake for humans), and toxicological aspects have been addressed.
Analysis of THg in muscles and liver of some pelagic, demersal and benthic
species captured inside and outside the semi-enclosed area, has been
analysed in order to explore the effects of HgT pollution on fish
compartment and to assess the potential health risks associated with the
consumption of contaminated fish. THg content of fishes shows a wide
range of concentration (range: 0.02 - 2.71 μg g-1
and 0.03 -9.72 μg g-1
in
muscles and in livers respectively), with highest values measured in benthic
species and the lowest in pelagic ones. This increasing trend along the
habitat depth suggests an active release mechanism of mercury from
polluted sediments to the water column, with consequent effects of
bioaccumulation in the trophic web. Anomalous THg content measured in
pelagic species captured in the external zone of the bay confirms the role of
the Augusta marine environment as pollutant source of Hg for the
surrounding area and underscores the crucial risk associated with
contaminant transfer from the basin to the open sea. Finally, values of
hazard target quotient (THQ) and estimated weekly intake (EWI)
demonstrate that consumption of fishes caught inside the bay represents a
serious risk for human health and suggests caution in consuming demersal
and benthic fishes from outside the Augusta Bay, definitively demanding for
appropriate social actions.
Hg distribution in sediments (range: 1.77 - 55.34 mgKg-1
; mean: 13.78
±10.72 mgKg-1
) clearly divides the area into three parts, with the lowest
values recorded in the northern Augusta Bay, intermediate value in the
center and the highest HgT concentrations recorded in the southern part of
the Augusta basin (from the Pontile Cementeria down to the dam) with
decreasing values from the coastline. Despite sequential extraction
procedure (SEP) documented that the most part of HgT in sediments
consists of strong complexes (~80% of HgT as strong complex, ~15% of
Hg as less strong forms and ~ 2% of HgT as more soluble and bioavailable
forms), some anaerobic microorganisms can manage these stable Hg
trapped in minerals structures as substrate for their metabolism, making Hg
more easily bioavailable for the environment. Analysis of THg and DHg
content at different quote of the water column, provided significant
information on Hg distribution along space and depth. The HgT distribution
in seawaters (range: 0.45 - 129.27 ngL-1
and >0.01 - 21.3 ngL-1
for THg and
Abstract
4
HgD respectively), putted in light an evident increasing trend of Hg content
toward the southern and more contaminated part of the Augusta Bay, where
waste spillage from chlor-alkali plant occurred. A clear trend was also
observed on the vertical, with Hg concentration increasing near the bottom
and reducing in surface water, strengthening the role of Augusta sediments
as sources for the overlying water. Moreover the unexpected THg
concentrations measured in seawater outside the bay (range: 2.62-11.95 ng
L-1
; mean: 6.46±2.95 ngL-1
), confirmed the hypothesis of transport of Hg
from Augusta harbor to the open sea representing a vehicle of contamination
for the entire Mediterranean basin through the complex circulating currents
affecting the western Ionian.
In this scenario, fluxes assessment at the interface sediments-seawater-air
became crucial in order to create a mass balance of Hg in the study area and
to determine the net outflow for the Mediterranean sea. Detailed information
on the mobilization processes from sediment to seawater and consequent
escape to the atmosphere has been investigated. For this reason, for the first
time in this area, a benthic chamber and a dynamic accumulation chamber,
have been employed in order to evaluate fluxes at the interfaces
sediments/seawater and seawater/air and to recognize equilibrium of
exchanges among phases. Using in situ accumulation chamber Bagnato et
al., 2013 reported an estimated sea–air Hg evasion for the entire Augusta
basin (~23.5 km2) of about 9.7 ± 0.1 g d
-1 (~0.004 t yr
-1), accounting for
~0.0002% of the global Hg oceanic evasion (2000 t yr-1
).
Simultaneously using in situ benthic chamber, a total flow from sediment to
seawater for the whole Augusta Bay has been estimated in 0.22 kmol y-1
in
2011 (0.05 ty-1
) and 0.38 kmol y-1
in 2012 (0.11 ty-1
). The mass balance
calculation permitted to estimate a HgT output from the Augusta basin to
Ionian surface waters (O) corresponding to an average of 1.29 kmol y-1
.
Analysis of Hg isotopes in sediments and fishes of the Augusta Bay,
provided unique information on Hg sources in the environment and
processes influencing Hg cycling. The success of such an approach strongly
depends on two factors. First, different natural and anthropogenic Hg
sources must have analytically discernible Hg isotope signatures. Second,
the processes that transport and transform emitted or discharged Hg into the
environment must not obscure the original Hg source isotope signatures.
This requires that fractionation of Hg isotopes after release is either small
relative to source differences or is predictable enough to be corrected for,
allowing estimation of the source isotopic composition. The magnitude of
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
5
mass-dependent (MDF) and mass-independent fractionation (MIF) has been
described primarily as δ 202
Hg and Δ201
Hg. The positive MIF fractionation in
fishes, especially in pelagic one, demonstrated photochemical reaction of
Hg(II) prior of the intake in the marine food web. Sediments isotopes
fractionation demonstrated reaction of methylmercury production biological
mediated. A geographic pattern in δ202
Hg and Δ199
Hg values suggests that
the sources of Hg to the sediment are locally controlled. Hairs exhibit
positive MIF fractionation, suggesting reaction of photochemical reduction
of MeHg in presence of organic matter. The overlapping δ 202
Hg values of
both sediments and fishes suggested sediments represent the source of Hg
for fish. The positive relationship obtained by plotting Δ201
Hg vs. Δ199
Hg of
both hairs and sediments demonstrate fish consumption represents the first
pathway of exposure for human. Difference of 2‰ between δ 202
Hg in fishes
and values, could be due to could suggest that substantial MDF takes place
during MMHg human metabolism
Rare Earth Elements (REEs) are important because their geochemical
properties enable them to be powerful tracers of chemical processes. Their
distribution has been investigated in seawater of the Augusta Bay in order to
verify if anthropogenic sign can also transpire through the investigation of
REE. The REEs distribution along the water column suggests that the high
dissolved organic matter created ideal condition for an increasing of REE in
dissolved phases, much to hide the negative Ce anomaly usually recorded in
the oligothropic water. Gd anomaly, expressed as Gd/Gd*>1, suggests
significant contributions of the petrochemical industries, using gadolinium,
in the form of gadolinium oxide, as petroleum cracking catalyst.
A common thread, started from the evaluation of Hg in the key component
of the cycle, the study of fluxes at the interfaces, the evaluation of Hg
isotopic fractioning and the REE distribution in water column, permitted to
evaluate the fate of Hg in the Augusta Bay and the main processes rule the
Hg biogeochemical cycle.
Abstarct
Abstract
6
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
Chapter I:
Introduction
7
1 Preface
The knowledge of the Hg complex biogeochemical cycle improves at the
beginning of the twenty-first century, when the potential risks of human
exposure to mercury (Hg), especially in the form of monomethylmercury
(MMHg) and the potential deleterious ecological consequences from
mercury pollution have given much impetus to mercury studies and
regulatory activities. Nowadays the growth in mercury research continues at
breakneck pace.
Studies of the mercury behavior in strongly impacted sites, especially in
coastal marine environment, result crucial in order to a deeper
understanding on the biogeochemical cycle of this dangerous element. In
this view the Augusta Bay, due to the careless discharge of Hg from
industrial and petrochemical plants, and the complex water system
involving the Sicilian coast, represents an ideal natural laboratory for
advancement on Hg geochemistry.
In particular, the southernmost part of the Augusta basin hosted one of the
most important chlor-alkali plants of Italy (Syndial Priolo Gargallo), which
made up over 20% of total Italian emissions in 2001 (Le Donne and Ciafani,
2008). In order to supply helpful data on the Hg behavior in a polluted and
complex coastal marine system, an integrated model on the biogeochemical
cycle of Hg in the Augusta Bay will be detailed.
The aims of this thesis are specified below.
1.1 Goals of the thesis
1.1.1 General goal
The main aim of this PhD thesis is the generation of integrated
biogeochemical models of mercury in the Augusta Bay by the analysis of
processes at the sediment-seawater-atmosphere interfaces and the transfer
in the trophic web
1.1.2 Specific goals
Study of Hg bioaccumulation processes in the biotic compartment
and risk analysis for human health;
Chapter I: Introduction
8
Analysis of Hg fluxes at the interfaces (sediment-water, water-
atmosphere);
Analysis of the metal speciation in sediments;
Mass balance calculation;
Study on Hg isotopes finalized to the assessment of sources and
pathways of mercury contamination;
Study of REEs distribution in the water column as tracers pollution.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
9
1.2 Mercury element
Mercury is a natural element of the earth, a lucid liquid metal, present in a
variety of chemical forms in rocks, soil, water, air, plants and animals. It is
found usually combined with both inorganic (e.g., the mineral cinnabar, a
combination of mercury and sulfur) or organic (e.g., MeHg) compounds
(EPA, 1980; Clarkson et al., 1984), although occasionally Hg occurs in its
elemental form, as (Hg0) in liquid or vapor forms.
Hg has not known biological function and it is potentially hazardous for
human health. Biological processes can convert forms of mercury with
relatively low toxicity, in the most toxic form, methylmercury (MeHg)
present in fish and seafood products, while in other foods, mercury is in
inorganic and not toxic form (WHO, 1991b).
Moreover mercury can bioconcentrate in organisms and undergoes
biomagnified processes through the food chains (e.g., Westöö, 1966;
Fitzgerald and Watras, 1989; Wiener et al., 1990; Gilmour and Henry 1991;
Watras and Bloom, 1994; Watras et al., 1994; Rolfhus and Fitzgerald, 1995;
Hall et al., 1998; Benoit et al., 2003).
Increasing of Hg in environmental compartment has been demonstrated to
have increased up to 5X pre-cultural levels, primarily as a result of man's
activities, sometimes with dramatic consequences on environment and
human health (Bakir et al., 1973; Smith, 1975; George, 2001).
1.3 Chemical properties
Mercury element (Hg, from the Latin hydrargyrum or “watery silver”;
Fitzgerald and Clarkson, 1991) is readily familiar as a silvery liquid at room
temperature and its highly volatility.
Mercury is the metal with the lowest melting point and with a vapor
pressure lower than liquids. Due to its high reduction potential, mercury
behaves in similar way to a noble metal at room temperature, resulting not
attacked by non-oxidizing acids, alkali, oxygen, nitrogen, phosphorus,
carbon, ammonia, hydrofluoric acid and hydrochloric acid, while it is rather
sensitive to nitric acid, bromide and iodide, halogens and ozone. At room
temperature mercury oxidizes very slowly, but its oxidation proceeds with a
useful speed to 300-350 °C.
Chapter I: Introduction
10
Mercury easily forms alloys, or amalgams, with almost all of the most
common metals, including gold and silver. Chemical speciation is probably
the most important variable influencing ecotoxicology of Hg (Boudou and
Ribeyre, 1983). Mercury can exist in three oxidation states: elemental
mercury (Hgo), mercurous ion (Hg
I), and mercuric ion (Hg
II). The oxidation
state +1 is presented in the form Hg22+
, in which the two atoms of mercury
are linked together with covalent bonds.
Mercury compounds in an aqueous solution are chemically complex and
influenced by a variety of factors: pH, alkalinity, redox, organic matter and
other variables. A wide variety of chemical species are liable to be formed,
having different electrical charges and solubility. For example, HgCl2 in
solution can speciate into Hg(OH)2, Hg2+
, HgCl+, Hg(OH)
-, HgCl
3-, and
HgCl42-
, anionic forms predominate in saline environments (Boudou and
Ribeyre, 1983).
In solution mercurous ion is stable only in excess of metallic mercury,
otherwise shift in the metallic mercury and mercuric ion. The chemical
forms of mercury of interest in biological and environmental samples are
listed in Table 1:
Table 1: Mercury compounds in environmental and biological samples.
Elemental mercury Hg (0)
Inorganic Hg forms mercuric ion Hg2+
Mercurous ion Hg+
Mercuric sulfide (o cinnabar) HgS
Organic Hg forms Methilmercury CH3Hg+
Ethylmercury C2H5Hg+
Phenylmercury C6H5Hg+
Dimethylmercury (CH3)2Hg
1.4 Sources and application
The remarkable Hg properties and its major mineralized form (cinnabar,
HgS) have been well known (Goldwater, 1972; Fitzgerald and Clarkson,
1991). It was initially used in decorations, cosmetics and religious rites, but
the introduction of mercury in science occurred in 1643 with the invention
of Torricelli's barometer, for atmospheric pressure measures, and in 1720
with the Fahrenheit mercury thermometer.
Other employs include the use of Hg as “spring tonics,” as a “cure” for
syphilis, and as panacea for other afflictions (Fitzgerald and Clarkson,
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
11
1991). Mercury can reach the environment through both natural and
anthropogenic source (Turekian and Wedepohl, 1961; Mason et al., 1994; ).
The first one includes volcanoes, both volcanoes aerial and sub-marine
activities (Patterson and Settle, 1987; Fitzgerald, 1981, 1996; Mason et al.,
1994), weathering inputs from mineralized mercury deposits (Koski et al.,
1994; Ozerova, 1996; Stoffers et al.,1999), low-temperature volatilization
(Xiao et al., 1991; Carpi and Lindberg, 1998; Gustin et al., 1999; Boudala et
al., 2000) and volatilization from the ocean (Fitzgerald and Clarkson,
1991). Various and widespread are the human-related sources of mercury to
the environment, and contribute to ~2/3 of the mercury emitted from land-
based sources each year (Pacyna and Pacyna, 2002).
In particular, inputs of mercury to the environment via the atmosphere,
particularly high temperature processes due to coal and municipal waste
burning, are of the greatest concern (NAS, 1978; Wang et al., 2000; Long
and Kelly, 2002; Pacyna and Pacyna, 2002; Wilhelm, 2001; Fitzgerald and
Clarkson, 1991). A short summary on global anthropogenic emissions of Hg
calculated by Pacyna et al., (2006) is reported in Table 1.
Table 2: Global anthropogenic emissions of Hg in 2000 (in ton) (Pacyna et al., 2006)
Continent Stationary
combustion
Cement
prod.
Non-
ferrous
metal
prod.
Pig
iron
&
steel
prod.
Caustic
soda
prod.
Mercury
prod.
Gold
prod.
Waste
disposal Other Total
Africa 205.2 5.3 7.9 0.4 0.3 0.1 177.8 1.4 398.4
Asia (Excl.
Russia) 878.7 89.9 87.6 11.6 30.7 0.1 47.2 32.6 0.9 1179.3
Australasia 112.6 0.8 4.4 0.3 0.7 7.7 0.1 126.6
Europe (Excl.
Russia) 88.8 26.5 10 10.6 12.4 11.5 15.3 171.1
Russia 26.5 3.7 6.9 2.7 8 3.1 3.5 18.2 72.6
South
America 31 6.5 25.4 1.4 5 22.8 92.1
North
America 79.6 7.7 6.4 4.3 8 0.1 12.2 18.7 8.8 145.8
Total 1422.4 140.4 148.6 31.3 65.1 23.1 248 66.4 44.6 2189.9
The category defined as ‘‘other sources’’ includes chlor-alkali production
using the mercury cell process and production of battery, electrical lighting,
wiring devices, switches and instruments (Pacyna et al., 2006). Although the
use of the mercury cell to produce caustic soda in the chlor-alkali industry
Chapter I: Introduction
Chapter I: Introduction
12
has decreased significantly over the past 10–15 years worldwide
(www.eurochlor.org), the global consumption of mercury in this industry
was still of ~800 ton in the year 2003 (Maxon, 2005).
Five chlor-alkali plant are still working in Italy: Tessenderlo of Pieve
Vergonte (VCO); Caffaro of Torviscosa (Ud); Syndial of Porto Marghera
(Ve); Solvay of Rosignano (Li); Solvay of Bussi sul Tirino (Pe). The sixth
one, Syndial of Priolo Gargallo (Sr), located in the Augusta bay, is closed
since 2005 (Le Donne and Ciafani, 2008).
Although some industries have established programs aiming at the reduction
of mercury releases to the environment, as the chlor-alkali industry
(EuroChlor, www.eurochlor.org), the phaseout of the mercury process in the
chlor-alkali industry will be prolonged after 2020 (Pacyna et al., 2006).
1.5 The biogeochemical cycle of mercury in the
environment
In marine environments, Hg undergoes a complex environmental cycle
occurring between sediments, water column and atmosphere, and that
implies the involvement of biotic compartment.
The balance between the reduced and oxidized Hg form depends
essentially on the redox reactions of these media. In this view sediments
can represent source of sink for Hg.
Inorganic mercury, present in environmental compartments, is mainly in
oxidation state +2. Due to its nature of sulfur loving (i.e., chalcophilic)
(Fitzgerald and Clarkson, 1991), the tendency of Hg2+
to bind with the
sulfur-groups is notable, especially where the concentration of HS- and
S2-
is high, as in anoxic waters and sediments, where mercury sulfide,
solid and slightly soluble in water, tends to accumulate into the sediments
(Langer et al., 2001; Gilmour and Henry, 1991).
Methylation of inorganic mercury is the most important transformation
affecting the behavior and fate of mercury in aquatic systems (Fitzgerald
and Clarkson, 1991).
Indeed MMHg production represents the main vehicle of transport from
sediment to seawater and fishes. In situ methylation of “reactive” or
bioavailable mercury in aquatic systems appears to be predominately
biotic (Benoit et al., 2003, Gårdfeldt et al., 2003) and causes MMHg
accumulation in freshwater foodwebs (e.g., Westöö, 1966; Wiener et al.,
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
13
1990; Gilmour and Henry 1991; Watras and Bloom, 1994; Watras et al.,
1994; Hall et al., 1998).
Benoit et al. (1999 a, b, 2001 a, b) proposed that sulfide-oxidizing
bacteria (SOB), being chemolithotrophs, are able to use sulfide as a
source of energy and reducing power.
The mechanism for uptake of inorganic mercury by methylating bacteria
is not known, Benoit et al., (2003) propose the mechanism of diffusion of
HgS0 through the cellular membrane as the key factor. Bacterial
mediation enhances the rates at which mercury, a “soft acid,” can form
alkylated species in aquatic environments.
Microbial production of MMHg in sediment is influenced the activity of
methylating organisms (i.e., SRB) and by the availability of inorganic
mercury for methylation. Maximum mercury methylation occurs in
sediments where organic matter and sulfate are sufficiently high as to
stimulate SRB metabolism, but not so high as to cause accumulation of
sulfide, reducing the availability of mercury for methylation (Gilmour
and Henry, 1991).
Other variables, such as dissolved organic carbon (DOC) and pH strongly
affect the ultimate fate of mercury in an ecosystem. An increasing water
acidity (decreasing pH) and/or DOC content, usually result in higher
mercury levels in fishes and in higher methylation rate.
Methylation amplifies the insidiousness of historic mercury pollution and
health risks to wildlife and humans. As well as inorganic mercury,
MMHg, has a high affinity towards the ions Cl-, Br
-, HS
and OH
and,
therefore, in environmental compartments is present in associated form in
different species depending on the ions concentration.
In natural waters MMHg levels are generally much lower than those of
inorganic Hg, due to the difficulty with which the methylation reactions
occur in aqueous phase and because of partial decomposition of mercury
due to sunlight. On the contrary in sediment and biota (especially in fish)
methylmercury concentration is much greater due to bioaccumulation
phenomena.
Although MMHg represents the dominant chemical form in most of the
organisms of the upper food pyramid, it constitutes only a small fraction
of the total mercury present in an aquatic ecosystem (Lindqvist et al.,
1984). Indeed once in the water column, in addition to the methylation
process, Hg2+
can undergo reduction processes leading to the formation
of volatile chemical species (Hg0) and return to the atmosphere though
Chapter I: Introduction
14
(Fig. 1) through which the ocean prolongs the residence of mercury at the
Earth’s surface.
The volatile chemical species of dissolved mercury are indicated by the
acronym DGM (Dissolved Gaseous Mercury) of which the elemental
mercury represents more than 90% (~1.25 yr; Hudson et al., 1995,
Fitzgerald and Gill, 1979; Bloom N. S. and Fitzgerald, 1988; Iverfeldt,
1991a), although some chemical forms of Hg2+
also exist in the gas phase
(the so-called “reactive gaseous mercury” or RGM; Stratton W. J. and
Lindberg, 1995; Sheu and Mason, 2001; Landis et al., 2002).
The reduction mechanisms of Hg2+
are due to photo-chemical reactions
(abiotic processes) and to the intervention of prokaryotic and eukaryotic
organisms (biotic processes) (Fig. 1).
Atmospheric oxidation of elemental mercury in atmosphere goes back to
the water column through dry and wet deposition (Lindqvist et al., 1991)
(Fig. 1). Mason, Fitzgerald, and Morel in 1994 (“MFM”) provided a Hg
biogeochemical model, suggested that about half of the modern pollution-
related emissions (20 Mmol yr-1
) enter the global cycle. Pacyna and
Pacyna (2002) reported comparable value, estimating for 1995 an
amount of ~9.6 Mmol yr-1
.
Figure 1: Scheme of the biogeochemical cycle of Hg in marine environmental system
(Fitzgerald and Clarkson, 1991 modified)
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
15
1.5.1 How does mercury enter the food chain?
The exact mechanisms by which mercury enters the food chain remain
largely unknown. As discussed above, certainly bacteria that process
sulfate (SO42-
) in the environment take up mercury in its inorganic form
and convert it into MMHg through metabolic processes.
The MMHg creation is important not only because its toxicity is greater
but also because organisms require times considerably longer to eliminate
it. Because animals accumulate methylmercury faster than they eliminate
it, animals consume higher concentrations of mercury at each successive
level of the food chain. Numerous biological and abiotic factors modify
the toxicity of Hg compounds, sometimes by an order of magnitude or
more.
Bellante et al., (2011) documented the existence of different mechanisms
of bioaccumulation, through different diet patterns and/or uniqueness in
physiological and/or biological control of Hg incorporation for S.
coeruleoalba and T. truncates of the Mediterranean sea. Different studies
demonstrated the MMHg accumulation in fish not only poses a threat to
human health (Grandjean et al., 1997; Davidson et al., 2000), but also can
reduce the reproductive success of piscivorous wildlife (e.g.,
Scheuhammer, 1991) and of fish themselves (Wiener et al., 1990b;
Wiener and Spry, 1996; Hammerschmidt et al., 2002).
The evidences of the Hg effects were assorted and demonstrated by an
extensive scientific literature, as well as news accounts of these tragedies.
Particularly, the Japanese poisonings resulted from consumption of
locally fish and seafood contaminated principally by MMHg abiotically
synthesizing as a by-product during the acetaldehyde production.
In the Iraqi tragedy, bread was the contamination source, which had been
made with flour, unknowingly milled from wheat treated with MMHg as
a fungicide (Bakir et al., 1973). The following works on the Japanese
poisonings provide an overview and useful starting point for a more in-
depth examination: Smith (1975) Japan Public Health Association
(2001), The Social Scientific Study Group on Minamata Disease (2001)
and George (2001). The USEFDA and the JECFA, 2000
recommendation is equivalent to 0.5 mg kg-1
of body weight per day. The
USEPA fix the reference dose in 0.1 μg Hg kg bw-1
d-1
(http://cfpub.epa.gov) while the acceptable daily intake determined by
WHO (2003) is 0.23 μg Hg kg bw-1
d-1
).
Chapter I: Introduction
16
1.6 Health concerns
The classification of mercury and its compounds, in accordance with
67/548/CEE Directive on the classification and labeling of hazardous
substances, is "toxic" and "dangerous in environment". The toxicity to
humans and other organisms depends on the chemical form, the amount, the
pathway of exposure, and the vulnerability of the exposed person (US EPA,
1989).
Human exposure to mercury can result from a variety of pathways,
including consumption of fish and fish products (primary pathway of
exposure), occupational and household uses, dental amalgams, and mercury-
containing vaccines. Mercury toxicity depends of the absence of
homeostatic mechanism that regulates and limits its accumulation in the
body. Mercury binds strongly with sulfhydryl groups, and has many
potential target sites during embryogenesis; phenylmercury and
methylmercury compounds (mono and dimethylmercury MMHg, DMHg)
are among the strongest known inhibitors of cell division. MMHg and other
organomercury compounds, cross the placental barriers and can enter
mammals by way of respiratory tract, gastrointestinal tract, skin, or mucous
membranes (Elhassani, 1983).
The toxicological properties of mercury for environment and human health
strongly depend on physical and chemical form in which it occurs. Hg
vapors for example, are very dangerous if inhaled, reach the lungs causing,
pulmonary edema, pain and peeling of the respiratory epithelium of the
bronchi.
Acute intoxication by organic mercury affects the central nervous system
causing tremors, memory problems and neurodegenerative diseases such as
Alzheimer's and Parkinson's. However only 10% of inorganic mercury is
ingested it is absorbed as salt of Hg2 and Hg
+, reach the human body
through the skin and the oral cavity, causing diseases related to the digestive
system and kidneys and renal toxicity in which cause apoptosis increment
(type of cell death, different from necrosis) for inactivation of nuclear
factor-kB (NF-kB), a thiol-dependent transcriptional system. Thiol groups
(R-SH), which compete with Hg for protein binding sites, are the most
important antagonists of inorganic mercury salts (Das et al., 1982).
Mercury-antagonistic drugs include 2,3-dimercaptopropanol, polythiol
resins, selenium salts, vitamin E, and sulfhydral agents (Nriagu et al., 1979;
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
Chapter I: Introduction
17
Magos and Webb, 1979; Elhassani,1983). The protective action of selenium
(Se) against adverse or lethal effects has been reported for algae, aquatic
invertebrates, fish, and mammals (Bellante et al., 2011; Magos and Webb,
1979; Heisinger, 1979; Chang et al., 1981; Lawrence and Holoka, 1981;
Das et al., 1982; Gotsis, 1982; Eisler, 1985; Satoh et al., 1985).
1.7 Mercury in Mediterranean Sea
1.7.1 The Mediterranean geochemical anomaly
The Mediterranean basin is characterized of relevant Hg sources of both
natural and anthropogenic origin (Bacci, 1989). Table 3 reports range or
mean mercury concentrations in intermediate and deep waters at different
stations of the Mediterranean sea and of the world ocean. The dispute on
what is the primary source that determines the so-called “Mediterranean
geochemical anomaly” is still active. The simultaneous presence of cinnabar
deposits, volcanoes and anthropogenic influences are the main causes. On
the one hand about 60% of the mercury deposits around the world is located
in the Mediterranean basin (Ferrara et al., 2000) (e.g. Almadén (Spain),
Idrjia (Slovenia) and Monte Amiata (Italy)), from where mercury has been
released by both natural and anthropogenic activities (Fig. 2).
Figure 2: Map of the main cinnabar deposits in Mediterranean basin (Ferrara et al., 2000
modified);
On the other hand mercury emissions into the atmosphere by the Stromboli,
Vulcano and Etna volcanoes, indicate volcanoes are a major source of
mercury in the Mediterranean (Bagnato et al., 2007, 2011). Using in situ
Introduction
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
Chapter I: Introduction
18
chambers, Covelli et al., 1999 determined mercury benthic fluxes of HgT
and MMHg in the Gulf of Trieste (Northern Adriatic), and reported a
mercury flux from exoreic runoff to the open Mediterranean of around 10
kmol yr–1
, ascribing to the shelf sediment a significant methylmercury
source for the ecosystems.
Moreover, the specific geological and hydrogeological features (poor water
exchange), together with the high solar radiation intensity and temperatures
that characterize the Mediterranean for many months of the year, promote
the biological, photo-chemical and physicists mechanisms responsible of the
formation and emission from soils and waters of elemental mercury (Ferrara
et al., 1986).
Particularly Bacci (1989) suggested that the higher temperature of the deep
Mediterranean waters might promote the bacterial methylmercury formation
in the water and/or sediments, with resulting higher mercury
bioaccumulation in fish tissues. Research conducted in recent years, have
estimated the total mercury emission from the water surface of the
Mediterranean Sea amounts to 60 tons year-1
, representing the primary
natural source of this metal to the atmosphere (Ferrara et al., 2000), higher
than that of volcanoes sources (0.6-1.3 tons year-1
, Ferrara et al., 2000).
In conclusion, while the Mediterranean mercury “anomaly” or specificity
was purported to be of a geochemical origin, the anomaly has probably
biochemical or ecological reasons, including the high methylating potential.
The Ionian Sea is one of the most oligotrophic zones of the Mediterrenean
sea and of the whole World Ocean. In summer 2000, the HgT
concentrations measured in water between the surface and 2500m varied
widely between 1.0 and 2.6 pM, and the methylated species were abundant
throughout the water column (10–26% of the HgT) (Horvat et al., 2003).
Table 3: Range or mean mercury concentrations in intermediate and deep waters at
different stations of the world ocean. HgT: total mercury; MeHg: methylated mercury
including MMHg and DMHg
HgT (pM) MeHg
(fM) MeHg/HgT (%) Reference
South and Equatorial Atlantic 0.8-2.4 25.200 ~7.5 a
North Atlantic 2.4±1.6 29-160 ~6.6 b North Pacific 1.2±0.9 <50 <4 c
Equatorial Pacific - 35-670 - d
Mediterranean
0.5-4.0 20-290 1-35 e
0.5-2.1 190-390 17-33 f
0.5-3.2 100-460 - g NW Mediterranean 0.8-1.2 Up to 210 Up to18 h
Chapter I: Introduction
19
a) Mason and Sullivan, 1999; b) Mason et al., 1998; c) Laurier et. al., 2004; d) Mason and Fitzgerald,
1993; e) Cossa et al., 1997; f) Horvat et al., 2003; g) Tessier et. al., 2004; h) Cossa and Coquery,
2005.
1.7.2 The Mediterranean global circulation
From the circulation point of view, the Mediterranean Sea is a semi-enclose
basin, connected with the Atlantic Ocean through the Gibraltar Strait and
with the Black Sea trough the Dardanelles Strait (Robinson and Golnaraghi,
1994). The surface evaporation rate is not in equilibrium with precipitation
and the balance is achieved by flux of water entering through the Gibraltar
Strait. On the Sicilian cost, surface waters, called Modified Atlantic Waters
(MAW), enter from the Atlantic Ocean, while the Levantine Intermediate
Waters (LIW) formed in the eastern basin as a process of surface cooling of
water masses which undergoes a severe salt enrichment (Ovchinnikov,
1984; Malanotte- Rizzoli and Hecht, 1988). The northern part of the Sicily
Channel is crossed by MAW (Raffa F. and Hopkins T. S., 2004) which
flows eastward in the upper part of the water column, led by the Atlantic
Ionian Stream (AIS). These surface waters meander, induced mesoscale
cyclonic and anticyclonic gyres (Lermusiaux and Robinson, 2001; Béranger
et al., 2004). The lower part of the water column, between about 250 and
400 m, is occupied by LIW which enters from the Ionian Sea through the
passage south of Malta and flows westward, forming, in meeting the MAW,
the anticyclone gyre in the Ionian coast.
Chapter I: Introduction
20
Figure 3: Sub-basin scale and mesoscale circulation features in the eastern Mediterranean.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
21
Chapter II: The study area
Chapter II:
The study area
22
2 The study area
2.1 The Augusta Bay
Due to the crucial geographical position, the Augusta Bay (Sicily, Italy)
hosts one of the most important European port in terms of good traffic. This
features have promoted, since the early 60th
, the born of one of the largest
and most complex petrochemical district in Europe (Sprovieri et al., 2011).
Nowadays the Augusta Bay is a high environmental risk area due to the
uncontrolled chemical discharges that occurred before the Water Pollution
Control Law (Merli law, 1978) and frequent industrial accidents (Ausili et
al., 2008).
These features make the Augusta Bay an ideal natural lab for the study on
the biogeochemical cycle of mercury in a coastal marine environment.
Moreover the complex water circulations system and the closeness with the
steep continental slope (part of the Malta escarpment), make the area a
potential point source of mercury for the entire Mediterranean Sea, with a
consequent large scale effect of Hg pollution.
2.1.1 Geographical feature
The area, situated between Capo Santa Croce and Punta Magnisi (Lat.
37,21°N - Long. 15,23° E) is a wide natural bay which covers ~ 30 Km of
the eastern Sicilian coast (Azzaro, 1993) (Fig. 4).
In the past years, part of the bay has been walled up with breakwaters
forming a vast harbor basin communicating with the sea through two narrow
inlets: the Scirocco inlet, 300 m wide and 13 m deep, and the Levante inlet,
400 m wide and 40 m deep (Sprovieri et al., 2011). The bay, ~ 8 Km long
and 4 km wide, covers a surface area of 23.5 km2 with a mean depth of 14.9
m.
The area can be divided into three main zones: the Xifonio harbor (north
area, outside the bay), the Augusta harbor (internal area) and the Priolo sine
(southern area, outside the Bay).
For “Augusta Bay” is intended the internal part of the harbor, within the
northern, central and southern breakwaters. The basin is characterized by
three different circulation systems. Water circulation in the Levante inlet is
dominated by a northward flowing, with different mean speed at the surface
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
23
(18 cm s-1
) and at the bottom (7 cm s-1
) (ICRAM, 2005), while the southern
current at the Scirocco inlet flows parallel to the coast with moderate speeds
(8 and 4 cm s-1
at the surface and the bottom, respectively) and the northern
portion of the basin is scarcely affected by active currents (ICRAM, 2005).
A very narrow shelf divided the Bay to the slope, part of the Malta
escarpment (Scandone et al., 1981; Sprovieri et al., 2011). Owing to the low
water turnover, the intense human activities and industrialization, the
Augusta Bay is a complex area with a high state of degradation (Sciacca and
Fallico, 1978; De Domenico et al., 1994).
Figure 4: The Augusta Bay, eastern Sicily (ICRAM, 2008, modified).
2.1.2 Geological feature
From a geological point of view, the site is characterized by sedimentary
land of different origins from Cretaceous to Quaternary ages, interspersed
with volcanic rocks flows. The area corresponds to a graben formed
between the end of the Pliocene and the early Pleistocene, limited by host
Chapter II: The study area
24
of limestone and calcarenite stones. Sediments filled the graben are the
products of the weathering of cliffs, volcanic rocks and limestone.
The water body is made up of coarse sands and organogenic calcarenite
stones, while the substrate consists mainly of clays of variable thickness.
Sands and calcarenite stones are present along the entire gulf. There are no
"geological" reason for a local nature background of Hg in the area.
2.1.3 The Augusta harbor
Due to the strategic geographical position, the amplitude (more than
300.000 m3) and the optimal meteorological condition, the Augusta Bay,
homes one the most important harbor of the entire Mediterranean sea, with a
global ability of about one million of tons, of which 30 millions of liquid
good and 500.000 tons of cements (Fig. 5)
(http://www.portoaugusta.it/1/il_porto_industriale_2176574.html) .
The commercial harbor (~5 km from the city center) is located at NW to the
bay and is characterized by intensive activities of loading / unloading of dry
bulk goods including, chemicals, fertilizers, iron, sulfur pills, concrete,
marble, wood, mechanical parts and fittings (wind towers), clinker, basalt,
and containers storage. Together with a commercial function, the Augusta
harbor, since 1934, has been played an important military role because it
houses a military base of the Italian Navy.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
25
Figure 5: The Augusta harbor.
2.1.4 Industrialization history
The importance of the port and its functions, as well as strategic location,
aroused interest in the creation of the industrial center. At the end of the
Second World War (since 1949) the process of industrialization has begun
in Italy.
The south-eastern part of Sicily reveals the most suitable in the national
project of the industrial revival in the post-war period upon the De Gasperi
government. In few decades, in the triangle of land including the
municipalities of Augusta, Priolo and Melilli, on a coastline of about 15 km,
it was planted the largest petrochemical complex in Europe, consisting of
oil, petrochemical and chemical industries (Fig. 6).
The beginning of the industrialization process can be dated back to 1948
when Angelo Moratti, a young industrialist better known as the Inter
football club president, decided to build in Augusta a refinery, the
RA.SI.OM (Refinery Sicilian Mineral Oils) subsequently sold to the Esso
management.
The choice of Augusta was determined by some features: primarily the flat
nature and the strategic geographical position, on the Suez-Gibraltar rout,
were fulfilled by the highest traffic of crude oil from the Middle East and
Russia, in second time by the wide availability of low cost manpower, the
Chapter II: The study area
26
presence of a natural port and the ability to use the underground tanks and
the pier of the Navy, used during the Second World War.
The refinery began to produce in 1950, giving rise to the so-called
"Economic Miracle", consisting of an oil hub, a cement plant and a
petrochemical plant with a chlor-alkali plant doted of mercury cell. Today
arise on the bay several major companies (ENEL Augusta, Esso Italiana
S.r.l., Sasol Italy S.p.A, Maxcom Petroli S.p.A, Polimeri Europa S.p.A.,
Dow Italia S.r.l, Syndial S.p.A, Buzzi Unicem) (Fig. 6), a production system
that consisted of a "integrated-cycle” which produces aromatic compounds,
polymers and intermediates for plastics production, synthetic lubricants,
rubber, paints, detergents, pharmaceutical products.
The process of industrialization changed dramatically a territory that from
agriculture was forcibly converted to industrial, causing a radical shift in
environmental, social and economic terms.
The lush natural backdrop of the Augusta bay changed radically; olive trees,
almond and citrus trees, and the salt marshes along the coast, gave way to
sheds, steel columns and chimneys.
Esso docks
SASOL docks
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
27
ISAB docks
“Cementaria docks”
Figure 6: Photo gallery of plants and piers in the Augusta Bay (photos of F. Bulfamante).
2.1.5 The chlor-alkali plant in Augusta Bay
The plant, owned by Montedison from 1958 to 1991, passed first to
EniChem, then to Syndial and putted out of service in 2003. The chlor-alkali
plant produced chlorine and caustic soda by means of the electrolysis of a
sodium chloride aqueous solution in a cell with graphite anode and metallic
mercury cathode (Castner-Keller method, or “Solvay method”) (Fig. 7).
About 1 mm of liquid Hg flows on an inclined plan of iron which
constitutes the cathode, above which, at about half a centimeter, are
suspended the anode, constituted of graphite horizontal plates.
The sodium chloride electrolysis with the Hg cathode gives rise to a
“sodium amalgam” that, for subsequent oxidation with water, forms a
solution of sodium hydroxide. The flow of current develops chlorine gas at
the anode, while at the cathode occurs the reduction of sodium, which melts
in forming mercury amalgam.
Chapter II: The study area
28
It was obtained caustic soda, hydrogen and mercury, than recycled. The
electrolysis of 1.5 t of NaCl with the Castner-Keller method consumes about
2 kg of graphite and about 0.2 Kg of Hg cathode, producing 1 ton of NaOH
and 0.87 t of Cl2.
Na+ +xHg +e
- Na(Hg)x (Mercury amalgam)
Electrode processes of the stack:
(pole -) 2Na(Hg)x 2Na++ 2xHg + 2e
-
(pole +) 2H2O + 2 e- H2 +2OH
-
_______________________________________
global process: 2Na(Hg)x +2H2O 2(Na+OH
-) +H2 + 2xHg
Figure 7: Hg cell of the chlor-alkali plant.
During the process, therefore, each ton of soda leads to losses of mercury in
the air (in small quantities) and in the processed waters (in a more massive)
of the order of 200g. This Hg lost, in absence of efficient recovery systems,
can produce serious pollution (Bacci, 1989).
Treatment with the sulphides is an excellent method to remove the ionic
species of mercury; this method, however, is ineffective for the dominant
forms of Hg in the discharges of chlorine-soda (Hg0 and calomel). This limit
may be overcome by coupling the use of sulphides with the use of
polysulfides (Na2Sn) that allows, at pH> 9, the removal of 93% of Hg0. By
coupling both methods at a pH ranging between 9 and 11, it is possible to
obtain the removal of 99% of total Hg.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
29
Increasingly restrictive regulations on the environment have led, in most
advanced countries, to the complete replacement of mercury cells with a
"membrane cell".
In Priolo replacement never happened, but the graphite anode of the chlor-
alkali plant was replaced in 1979 and steel electrodes were introduced and
the chlor-alkali was equipped with of a special system through which the
waste water is first treated with thiourea, able to precipitate mercury in the
very low solubility form of mercuric sulphide, than clarificated and
filtratered on sand and activated carbon.
The chlor-alkali plant discharged through the “Vallone della neve”
(southern Augusta bay), in amount variable in time. Mercury that causes
contamination of sediments of the Augusta Bay derives from untreated
wastewater of the chlor-alkali plant (1958-1979).
About 500 tons of mercury were spilled to the sea without treatment
between 1958 and 1979; negligible is the next quantity.
2.2 Human health and environmental concerns in
Augusta Bay
The first evidences of pollution effects on environment appeared in the early
80th
, when the Muscatello hospital (Syracuse) denounced an increase of
malformed born and the entire area was included into the I.P.I.M.C
monitoring program on congenital malformations.
The main evidence regards defects birth, heart and circulation, digestive
system, uro-genital system (hypospadias) cancer and rates abortion double
respect of the rest of the province and four times higher respect of national
reference (ASMAC, 1990; Madeddu, 2006; WHO, 2001, ENEA, 2001 ASL
8 of Syracuse, Epidemiological Observatory of the Health of Sicily, 2005).
In November 1990 the industrial area between Syracuse and Augusta was
declared to "high risk of environmental crisis”.
2.3 Environmental Hg pollution in Augusta Bay
Together with data on human health, environmental degradation of Augusta
Bay was well documented. Indeed in the last decade, several studies
Chapter II: The study area
30
provided detailed information on the pollution levels and risks for human
health of resident populations of Augusta Bay (ICRAM, 2005; Ausili et al.,
2008; Di Leonardo et al., 2007, 2008; ENVIRON International Team, 2008;
Ficco et al., 2009; Sprovieri et al., 2011).
Data recently collected by ICRAM (2008), ENVIRON International Team
(2008) Ausili et al. (2008), demonstrated HgT transfer from the abiotic
system (sediments and seawater) to fishes (top predators and filter-feeders)
and documented significant health risks associated with the consumption of
fish caught in the area.
Toxicological Hg effects were also evaluated on mussels and red mullet by
micronuclei (MN) studies, documenting DNA damages (Ausili et al., 2008
and ICRAM, 2008), in agreement with Tomasello et al. (2012) reported on
DNA genotoxic and oxidative damages in Coris julis specimens from
Augusta Bay. Fantozzi et al., (2012) found high Hg content in seawater
collected close to Augusta.
Sprovieri et al., in 2011 speculated on the key role played by the Augusta
bay as Hg point source for the surrounding marina environment and
reported high-resolution maps of HgT distribution from superficial
sediments collected in 2005, highlighting the extremely high concentrations
of Hg in sediments (Fig. 8).
With the exception of Sprovieri, ICRAM and ENVIRON investigations, no
more data are available on Hg distribution inside the bay.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
31
Figure 8: High resolution map of Hg distribution in sediments reported
(from Sprovieri et al., 2011).
Chapter II: The study area
32
Chapter III: Methodology
Chapter III:
Methodology
33
3 Methodology
This chapter details on the sampling strategy, treatment and analyses
finalized to Hg analyses in sediments, fishes, seawater and atmosphere. The
methodology used for flux assessing Hg at the interface between the
exchange surface (sediments-seawater and air-seawater) was also reported.
Analyses of REEs in seawater and of Hg isotopic analyses in sediments,
fishes and human hairs are also detailed.
Temperature data were acquired from the continuous acquisition by a
weather station (DAVIS – Vantage Pro 2Wi-Fi) installed on the roof of the
Augusta port authorities office.
3.1 Sampling strategy
Sampling activity has been carried out during different seasons and periods
of the years 2011-2012.
Details about sampling periods, kind and number of samples, relative station
sites and collection methods are reported in Tab. 4 and 5 and in Fig.s 9, 12
and 13. The sampling plan has been planned in function of the mercury
distribution in the Augusta Bay previously reported by Sprovieri et al.
(2011) (Fig. 8).
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
34
Figure 9: Sampling location of seawaters, sediments and benthic chambers station
planned in Augusta Bay.
Chapter III: Methodology
35
3.1.1 Sediment sampling
In May 2011 sediments were collected in station 7, 11 and 16 by means of
the research vessel “Dallaporta” in four station points of the Augusta Bay by
means of a box-corer instrument (Fig. 10a). Sediment collection during the
campaign of June 2012 (station 7, 12, 16bis) (Tab. 4; Fig. 10b), was carried
out by scuba divers with Plexiglas tubes (30cm long and a diameter of 6cm)
inserted into the sediments and carefully recovered and transported onboard.
All sediments were immediately sub-sampled onboard using a pre cleaned
(HNO3 10%) acrylic tube, sealed in polyethylene flasks and stored at −20 °C
until analysis.
Figure 10: Sampling sediment strategy; box-corer instrument (a) and manual sediment
sampling (b).
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
36
3.1.2 Seawater collection
Seawater samples are collected in May 2011 by means of the research
vessel “Dallaporta” in 15 stations of the Augusta Bay (Tab 4; Fig. 9) by
means of Niskin bottle rosette (Fig. 11a) at three different depth s (surface,
mid-water, bottom) at variable shares depending on the total bottom depth.
On February 2012 water samples has been collected in four points (Fig. 9,
station 18-21) outside the Augusta Bay at three depths on board of the
URANIA research vessel. Finally sample waters have been collected in June
2012 on board of the “Luigi Sanzo” vessel in three point (Fig. 9, station 7,
12 and 16) by means of a single bottle carried by a sub at the three depths of
interest (Fig. 11b).
In order to avoid samples contamination, ultra-trace Hg handling techniques
were employed during the collection and the analysis of water samples (US-
EPA, 2004). All containers were cleaned prior to use with HNO3 (10%) acid
and rinsed three times with Milli-Q water (N18.2 MΩ cm−1
). Samples were
stored in sterile bags at -20°C until the analysis were carried out (Covelli et
al., 2008; Horvat et al., 2003).
The Niskin rosette has been doted of a multiparameter probe able to acquire
environmental data (T, salinity, pH, eH).
Figure 11: Seawater sampling by means of a rosette Niskin bottles (a) and a single Niskin
bottle used manually by a scuba diver (b).
Chapter III: Methodology
37
Table 4: Sampling details with kind of sampling, employed vessel, sampling period, real
station positions and used instruments.
Sampling Vessel Sampling
period
Station
name
Station point Instrument
Lat. Long. S
eaw
ate
rs
N/O
G.
Dallaporta
23-26/05/11
1 37°14.392N 15°12.537E
Niskin
2 37°13.864N 15°12.519E
3 37°13.863N 15°11.845E
4 37°13.353N 15°11.902E
5 37°12.849N 15°12.595E
6 37°12.743N 15°13.176E
8 37°12.267N 15°11.863E
9 37°12.009N 15°12.161E
10 37°11.961N 15°12.855E
11 37°11.696N 15°12.512E
13 37°11.943N 15°13.536E
14 37°11.434N 15°13.536E
15 37°11.445N 15°12.865E
16 37°11.363N 15°12.434E
17 37°10.871N 15°12.833E
N/O
Urania 2/02/12
18 37°10.310N 15°
13.148E
Niskin
19 37°11.807N 15°
14.328E
20 37°11.742N 15°
15.767E
21 37°11.604N 15°
18.240E
M/N
L. Sanzo 23-26/06/12
7 37°12.578N 15°12.583E Single Niskin
bottle 12 37°11.697N 15°12.917E
16 37°11.288N 15°12.459E
Sed
imen
ts
N/O
G.
Dallaporta
23-26/05/11
7 37°12.618N 15°12.473E
Box-corer 11 37°11.818N 15°12.540E
12 37°11.745N 15°12.985E
16 37°11.399N 15°12.375E
M/N
L. Sanzo 23-26/06/12
7 37°12.578N 15°12.583E
Scuba diver 12 37°11.697N 15°12.917E
16 37°11.288N 15°12.459E
Sed
imen
ts-
sea
wa
ters
M/N
L. Sanzo 19-21/09/11
7 37°12.618N 15°12.473E
Benthic
chamber
11 37°11.818N 15°12.540E
16 37°11.399N 15°12.375E
M/N
L. Sanzo 23-26/06/12
7 37°12.578N 15°12.583E
12 37°11.697N 15°12.917E
16bis 37°11.288N 15°12.459E
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
38
3.1.3 Fish sampling
Four different sampling sites were selected: two inside, and two outside the
Augusta bay (Fig. 12). Sampling outside the bay was performed during May
2001, on board of the N/O “Dallaporta”, by means of a mid-water trawl-net
at 50-100 m of depth in two sampling areas, in front of the Scirocco inlet
(300 m wide and 13 m deep), and of the Levante inlet (400 m wide and 40
m deep) (Fig.12, points C1, C2). Mainly pelagic fish specimens were caught
(Tab. 5). Sampling inside the bay was performed during May 2012 by
means of a fishing boat equipped with a gillnet wall, positioned at the
bottom (mean depth = 20-25 m) (Fig 12, points C3, C4). Several specimens
of benthic and demersal fishes were collected. From the two sampling
activities, a total of 227 fish specimens were collected: 107 from mid-water
sampling (outside the bay) and 120 from bottom-water sampling (inside the
bay). Moreover, specimens of Engraulis encrasicolus (n = 38) were caught
from the unpolluted marine area of Marsala (western Sicily), during July
2001, on board of a fishing boat equipped with a purse seine net. After
collection, fishes were stored at T= -20°C until biological and chemical
analyses.
Table 5: Number of specimens per species caught in the sampling sites.
Mid-water sampling (outside the bay) Bottom-water sampling (inside the bay)
Species C1
(n°)
C2
(n°) Habitat Species
C3
(n°)
C4
(n°) Habitat
Engraulis
encrasicolus 20 20 Pelagic Diplodus annularis 59 15 Demersal
Sardina pilchardus 8 20 Pelagic Diplodus vulgaris - 3 Demersal
Boops boops - 20 Pelagic Pagellus erythrinus 1 6 Demersal
Trachurus trachurus - 6 Pelagic Pagellus acarne 11 1 Demersal
Illex coindetii 6 - Pelagic Sepia officinalis 2 6 Demersal
Loligo forbesi 3 - Pelagic Serranus scriba 2 - Demersal
Pagellus erythrinus 1 - Demersal Caranx rhonchus 1 - Pelagic
Pagellus bogaraveo 2 Demersal Sphyraena sphyraena 1 Pelagic
Mullus barbatus 1 - Benthic Scorpaena notata - 5 Benthic
Scorpaena scrofa 3 1 Benthic
Mullus barbatus - 3 Benthic
Mullus surmuletus 1 1 Benthic
Murena helena 1 - Benthic
Octopus vulgaris - 1 Benthic
Chapter III: Methodology
39
Figure 12: Sampling sites of fishes in the Augusta bay. The map reports also the
distribution of total mercury (HgT) in bottom sediments (Sprovieri et al., 2011
modified).
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
40
3.1.4 Atmospheric GEM assessment
In order to assess how elemental gaseous mercury GEM concentrates and
distributes in the Marine Boundary Layer (MBL) above the Augusta basin,
measurements of GEM were performed across the Augusta basin on-board
the Italian CNR research vessel Luigi Sanzo, during three main
oceanographic cruises carried out along the same route in the winter
(November 2011) and summer (July 2011-June 2012).
The cruise paths of the campaigns to date are shown in Fig. 13a. The
atmospheric investigation was finalized to assess the GEM content over the
land and the marine boundary layer (MBL).
The analysis of atmospheric GEM was performed using an automated real-
time atomic absorption spectrometer (Lumex-RA 915+). Wet and dry
deposition collection were been collected in the area. Details on atmospheric
GEM measurement and deposition collection were reported in Bagnato et
al., 2013.
Figure 13: Maps showing the (a) route of the cruises performed to measure GEM
concentrations in the MBL over the Augusta basin, (b) the land trajectories to detect
GEM contents in the atmosphere over the coastal internal area, (c) stations of
measurements of Hg evasion flux by accumulation chamber technique. Map (c)
Voronoi polygons, each relative to one station of measurement. (Bagnato et al., 2031)
Chapter III: Methodology
41
3.2 Flux at the interfaces
3.2.1 Fluxes at the sediment-seawater interface
In order to assess Hg fluxes at the interfaces among sediments, seawater and
atmosphere, an in situ benthic chamber and a dynamic accumulation
chamber were employed simultaneously in Augusta bay. The adopted
strategy to study fluxes of Hg through the interfaces is summarized in Fig.
14. The used in situ benthic chambers allowed to determine benthic fluxes
of dissolved mercury at the water–sediment interface throughout the time.
The benthic chamber was built by the technical staff working in the
laboratories of electronic at IAMC-CNR (Capo Granitola), according to the
different schemes proposed in literature (Covelli et al., 1999, 2008; Point et
al., 2007). The box-shape chamber (50x50x30cm), open at the bottom side,
was constructed from Plexiglas and equipped with a stirring mechanism
which consisted of a rotating bar (30 cm long; 5 rpm speed) inside the
chamber connected through a magnet to an electromotor, coupled with 12 V
batteries and located in a separate housing on the top of the chamber. A
plastic skirt, fitted to the outside of the chamber, controlled its penetration
into sediment to a depth of about 7.5 cm (Fig. 15).
Figure 14: Working scheme adopted in order to study fluxes at the interfaces.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
42
Figure 15: Benthic chamber built at the IAMC-CNR laboratories (Capo Granitola)
implied for the sediment-water sampling.
Benthic fluxes were assessed by deploying the in situ benthic chamber on
the sea-floor, isolating an area of sediment surface and overlying water.
Sampling at the sediment-water interface were carried out in Augusta Bay
on September 2011 (Station 7, 11, 16; Tab. 4) and June 2012 (Station 7, 12,
16bis; Tab. 4). The chamber was gently placed on the sea bottom by a scuba
diver and water samples from the benthic chamber were periodically
collected, at t=0, t= 1, t=4, t=6, t=10 and t=12 h, by scuba by means of a 50
ml syringes, inserted in a cap equipped with rubber septum pierceable (Fig.
16). Water was immediately transferred to the acid-pre-cleaned containers
and stored frozen.The sampling strategy adopted by sub has been detailed
by Patti et al., 2013 technical report.
Benthic fluxes of Hg dissolved species (DHg) were calculated from the
difference in concentrations, ΔC = Cf - C0, over the experiment time, Δt =
(tf - t0), following the Eq. (1):
t
AVCHg
/ (1)
where tf and t0 are the final and starting times, respectively, V is the benthic chamber
volume (57.5 L) and A is the sea-bottom area covered (0.25 m2) (Santschi et al., 1990).
Chapter III: Methodology
43
Figure 16: Sub aerial and submarine imagines of benthic chamber setting in the bottom
sediment by scuba diver (photo of V. Di Stefano and C. Patti).
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
44
3.2.2 Fluxes at the air-seawater interface
The open-bottom (50×50×50cm) dynamic flux chamber (made up of
plexiglass) used to estimate the flux during the cruises, was built by the
technical staff working in the laboratories of electronic at IAMC-CNR
(Capo Granitola), according to the schemes proposed in literature by Kim
and Lindberg, 1995 and Carpi and Lindberg, 1998.
The transparent plexiglas was chosen in order to permit solar radiation
penetration. The 125 liters flux chamber was used in conjunction with a RA-
915+ Lumex Hg vapor analyzer connected to the inlet and outlet of the flux
chamber situated about 20 cm above the water surface (Fig. 17). Details on
accumulation chamber experiments were reported by Bagnato et al., 2013
(Attachment II).
Fig. 17: Positioning and in real-time measurements of sea–air GEM evasion flux by using
the accumulation chamber technique.
Chapter III: Methodology
45
Mercury flux from the water surface exposed in the chamber was then
calculated to the Eq. (2) (Lindberg and Price, 1999; Zhang et al., 2001)
ФGEM= Q (C0-Ci)/A (2)
where GEM is the GEM total emission rate per area and unit time (ng m-2
sec-1
); (Co-Ci) is the difference in GEM concentrations in air exiting (Co)
and entering (Ci) the chamber (ΔC) (in ng m-3
); A is the basal area of the
chamber in m2; and Q is the flow rate of ambient air flowing through the
chamber in m3 sec
-1.
3.3 Samples preparation and analysis
3.3.1 Sediment treatment for THg analyses
Sediment samples were treated at sedimentology laboratory of IAMC-CNR
(Capo Granitola) (Fig. 18). Each core was defrost, extruded by liner and,
after macroscopic description, sectioned in cut at 1-3 cm intervals with a
stainless steel bandsaw. Aliquot of ~3.5 gr was used for grain size
composition, the remainder sediment slices were dried at 40 °C (Di
Leonardo et al., 2006) in order to remove water content.
Dry sediments were powdered manually in an agate mortar for analyses of
THg. Grain size analyses were performed following the procedure reported
by ICRAM, 2008.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
46
Figure 18: Sediments treatment at sedimentology laboratory of IAMC-CNR (Capo
Granitola).
Chapter III: Methodology
47
3.3.2 Sequential extraction procedure (Bloom et al., 2000)
Speciation is critical to understanding and modeling Hg contaminated soils
because the particular distribution of compounds and their interaction with
soil under aqueous conditions determine their environmental mobility and
bioavailability (Davis, et. al, 1997). To address this need, the robust solid
phase analytical scheme based upon sequential selective extractions of
Bloom, et al., 2000 was applied. This method does provide precise and
accurate data regarding the biogeochemically relevant fractionation of Hg in
sediments and soils. The method differentiates Hg into behavioral classes
including water soluble, ‘stomach acid’ soluble, organochelated, elemental,
and mercuric sulfide. Inorganic Hg speciation was determined by sequential
selective extractions of separate 2 gram aliquots of the homogenized solids
(Bloom, et. al., 2000).
The extraction was performed using a 100:1 liquid-to-solids ratio in 40 mL
vials. Sequential extraction procedure has been applied to May 2011
collected core sediments.
Each extraction step was conducted for 18±3 hours, with constant agitation,
at 18-22 °C. At the end of each step, the samples were centrifuged, and the
supernatant liquid was filtered through a 0.2 μm filter (Fig. 19). The solid
pellets were then re-suspended in the same extractant, re-centrifuged, and
re-filtered.
The two filtrates were combined in a 125 mL bottle, oxidized by the
addition of BrCl, and diluted to 125 mL prior to analysis for total Hg by
EPA Method 1631. After the rinse step, the sample pellet in the centrifuge
tube was resuspended in the next extractant, and the entire process was
repeated, according to the step specified in Table 6.
Table 6: Sequential Extraction Method Summary
Step Extractant Description Typical compounds
F1 DI water water soluble HgCl2, HgSO4
F2 pH 2 HCl/HOAc “stomach acid” HgO
F3 1N KOH organo complexed humics, Hg2Cl2
F4 12N HNO3 strong complexed mineral lattice, Hg2Cl2, Hgo
F5 aqua regia cinnabar HgS, m-HgS, HgSe, HgAu
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
48
Figure 19: Centrifuge, extractant and other tools used for SEP procedure.
3.3.3 Fishes treatment for THg analyses
The total length (TL) of each specimen was measured (Fig. 20). Muscle and
liver tissues were collected using plastic materials, cleaned with HNO3
(10%) and MillQ water, in order to avoid Hg contaminations. Tissues were
stored at -20°C until THg analysis.
Otoliths were extracted from anchovy and sardine specimens for age
determination. Readings and interpretations of otolith increment growths
were carried out according to the procedure adopted for European anchovy
age determination follow Uriarte et al. (2007) and La Mesa and Donato
(2009).
Chapter III: Methodology
49
Figure 20: Fish treatment for THg analyses
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
50
3.3.4 Seawater treatment for THg and DHg analysis
After the whole liter of water was defrost and shaking, aliquot of 50 ml was
treated with a BrCl solution (250 μl) for HgT analysis according to the EPA
7473 method. The remained part was filtered with polycarbonate filters
(Fig. 20) and treated with BrCl for HgD analysis. Samples were stored at
room temperature and analysed after 24h. KBr and KBrO3 are Hg free. 0.27
gr of KBr were dissolved in 25 ml of HCl and shaken for ~ ½ h, finally 0.38
gr of KBrO3 were added to solution that assumed yellow color. Chemical
reagents were suitable for ultra-trace Hg analysis. The THg and DHg
analysis in seawater, performed by means of the DMA-80 instrument.
3.3.5 THg analyses and data quality
Hg analyses in all the collected samples were performed using a direct
mercury analyzer (DMA80, atomic absorption spectrophotometer,
Milestone, Wesleyan University, Middletown, CT, USA) with optical path
spectrophotometer, achieving a detection limit of 0,0015 ng of mercury
(Fig. 21).
The DMA-80 is fully compliant with US EPA method 7473 (“Mercury in
solids and solutions by thermal decomposition, amalgamation, and atomic
absorption spectrophotometry”) and with ASTM method D-6722-01
(“Standard test method for total mercury in coal and coal combustion
residues by direct combustion analysis”). DMA-80 instrument uses the
principle of thermal decomposition, amalgamation and atomic absorption at
253,65 nm.
A total amount of 787 samples have been analysed by means of DMA-80
instrument, divided into: 462 fish tissues (231 muscles and 231 livers), 132
seawaters (66 for THg and 66 for DHg), 139 portion of sediments (85 for
THg and 54 for SEP procedure), 54 seawaters at the interface with sediment
(27 for THg and 27 for DHg).
Each sample has been analysed in duplicate or triplicate for a total number
of ~2000 analyses. In order to taste precision and accuracy in solid samples,
reference standard materials were measured each 10 measures.
Analyses of liquid (seawater, porewater) ~400μl of samples were placed
into quartz boats and analysed using a calibration range of 0-0.05 ng (Fig.
21). Hg liquid standard solution (Std 100 ngL-1
) was analysed in order to
Chapter III: Methodology
51
assess analytical accuracy. The detection limit, (Ld) and the critical value,
(Lc) for liquid solutions have been calculated following the “Water
Research Centre” (Cheeseman and Wilson 1979) procedure.
The critical value, Lc, ant the detection limit, Ld, (Tab. 7) were calculated
using the following equations:
wbLc 33.2 wbLd 65.4
where σwb is the within batch standard deviation of the blank (the term
“within-batch” signifies analyses made under the same experimental
conditions at essentially the same time).
For the instrument detection limits calculation, replied of blank sample has
been performed long time (n=10), resulting a instrumental detection limits
of 10.03, 20.02, 43.06 ngL-1
for Lc, Ld and Lq respectively. Results of the
blank measures were subtracted for the value calculation.
For HgT measurement in sediment samples, an aliquot of ~0.05 g was
loaded in nickel boats and transferred into the DMA-80 system (calibration
range: 5-500 ng).
Different standard materials were chosen and analysed, depending on the
sample Hg enrichment (Reference Standard Materials PACS-2 or MESS-3),
in order to test accuracy and precision (~7 and 6% respectively).
For HgT measurement in fish tissues, a wet aliquot of ~0.1 g was loaded in
nickel boats and transferred into the DMA-80 system (calibration range: 1-
200 ng).
A Reference Standard Material (TORT-2) was analysed to assess analytical
accuracy precision (4 and 7% respectively). Finally, duplicated samples
(about 20% of the total number of samples) were measured to estimate
reproducibility (~7%). Table 7 detailed reports data used for quality control
data and limits of detection.
Particularly, the precision, or variation coefficient, has been calculated by
standard deviation (σ) and mean value (Gini, 1912):
100*(%)Prmean
ecision
The accuracy (%), or relative error (%) has been calculating following the
ISO GUM (International Organization for Standardization, Guide
Uncertainty Measurement):
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
52
100*(%)
Y
YxA
Where X represent the measured value and Y the reference value.
Table 7: Quality data
Reference
material
N°
replies
Mean
THg
Measured
S.D. Accuracy
(%)
Precision
(%)
Sediments
Mess-3
(0.091±0.009
mg/Kg)
22 0.085 0.005 6 6
Pacs-2
(3.04±0.20
mg/Kg)
19 2.77 0.19 9 6
Fishes
Tort-2
(0.27±0.006
mg/Kg)
22 0.280 0.002 4 7
Seawaters Std (100 ng/L) 10 95.66 5.79 4 6
Limits (calculated on
the Blank)(ngL-1) Lc=10.03 Ld= 20.02
Figure 21: Direct Mercury Analyzer (DMA-80 Tricell). a) Principle of operation; b)
Related software; c) sample weight in a nikel boat; d) typical calibration.
Chapter III: Methodology
53
3.3.6 REEs treatment and analyses in seawater
Sample for REEs analyses were collected during June 2012 sampling (Fig.
9) in stations point 7, 12 and 16. Treatments and analyses were performed at
the University of Palermo following the pre-concentration technique
detailed by Raso et al. (2013). The typical features of the distribution of
shale-normalised REE concentrations, [REEi]n, calculated according to the
expressions:
(3)
with respect to Post Archean Australian Shales (PAAS; Censi et al., 2007),
By studying enrichments or depletions of single elements along the series,
usually named “anomalies”, the evaluation of “geochemical behaviour” of
REE is carried out. These anomalies can be assessed according to the
equation:
(4)
where the subscript “i” indicates every element along the REE series
whereas “(i-1)” and “(i+1)” are its immediate neighbour before and after
within the series (Alibo and Nozaki, 1999a). Only Gd anomaly can be
expressed according a different equation (Moller et al., 2007):
(5)
where the subscript “n” is referred to normalised concentrations. Features of
normalised-REE patterns can also be evaluated considering enrichments or
depletions of groups of REE subdivided into light REE, from La to Sm
(LREE), middle REE, from Eu to Dy or Ho (MREE) and heavier REE, from
Ho or Er and Lu (HREE) according to their atomic weight. These features
allow to define the geochemical behaviour of these elements during natural
processes.
11
*
2
ii
i
i
i
REEREE
REE
REE
REE
referencei
samplei
niREE
REEREE
3*
n
nn
Tb
HoGd
Gd
Gd
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
54
3.3.7 Hg isotopes analyses
Sediments, fishes and hair samples chosen for Hg isotopes analyses are
listened in table 8.
Table 8: List of fish, sediment and hair samples, with relative THg content (mg Kg-1
)
selected for Hg isotopes
Fish Sediments Hairs
Inside/
outside
THg
(mgKg-1) St. Depth
THg
(mgKg-1) Age
THg
(mgKg-1)
Sardina pilchardus O
Pel
ag
ic F
ish
es
0.04
Sta
tion
7
0-1 8.34 20 1.71
Sardina pilchardus O 0.04 2-3 6.79 30 1.34
Sardina pilchardus O 0.09 4-5 6.42 30 1.46
Sardina pilchardus O 0.11 13-
14 8.66 35 0.45
Engraulis encrasicolus O 0.04 14-
16 6.84 35 0.72
Engraulis encrasicolus O 0.04 18-
20 7.23 35 0.72
Boops boops O 0.13 24-
26 5.74 35 3.06
Boops boops O 0.16 26-
28 4.12 40 1.37
Trachurus trachurus O 0.31 30-
32 5.59 40 3.43
Trachurus trachurus O 0.33 36-
37 5.64 40 3.42
Sphyraena sphyraena I 2.27
Sta
tion
12
0-2 15.96 40 2.32
Diplodus annularis I
Dem
ersa
l
Fis
hes
1.42 2-4 13.36 40 1.49
Pagellus acarne I 0.26 6-8 12.16 40 1.47
Pagellus acarne I 0.70 8-10 16.83 40 1.83
Pagellus acarne I 0.26 10-
12 12.80 40 1.42
Pagellus acarne I 0.26 14-
16 9.33 40 5.07
Pagellus erythrinus I 0.42 18-
20 39.96 40 2.61
Pagellus erythrinus I 0.47 20-
22 8.01 40 1.97
Pagellus erythrinus I 0.35 22-
24 9.12 40 4.81
Mullus barbatus I
Ben
t
hic
fish
e
s
0.71
Sta
ti
on
16
0-2 17.11 40 5.28
Mullus barbatus I 0.71 4-6 18.99 40 2.74
Chapter III: Methodology
Table 8 continued
55
Mullus surmuletus I 0.60 10-
12 28.13 40 1.28
Scorphaena notata I 1.65 12-
14 38.41
14-
16 55.34
18-
20 39.96
24-
26 18.51
28-
30 22.08
34-
36 1.78
Samples for Hg isotopes analyses were achieved after total digestion. ~50
mg of dry and homogenized sediment were treated with a strong acid
mixture (9ml of HNO3 + 3 ml of HCl ) in a closed microwave system, while
6 ml of HNO3 were added to ~20 mg of dry and homogenized fishes
muscles prior the microwave digestion and hair samples were dissolved
with 2 HNO3 + 0,5 H2O2.
Isotopic analyses were performed at the UNIMORE Laboratory of Modena
by cold vapor multi-collector inductively coupled plasma mass spectrometer
MC-ICP-MS (Thermo-Fisher Neptune) (Fig. 22).
The standard bracketing technique (Fig. 23) has been used in order to
compensate the effects of mass bias changes during the measures sequences.
The technique consists of a specific acquisition sequence, in which two
standard measures are placed before and after each sample analysis. The
average standard values are used for δ calculation. Moreover each sample
or standard measure is immediately preceded and followed by two white
samples.
The two washing solution consisted of a solution of HNO3 4% in H2O
MilliQ® and of L- cysteine, able to form the complex Hg(Cys)2.
The intensity of the detected signal are subtracted, cup for cup, to those of
the respective standard or sample.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
56
Figure 22: Scheme of the ICP multicollector spectrometer Neptune (Berni PhD Thesys).
Figure 23: Block diagram of the readings and calculations sequence for RI determination
using the technique of bracketing standard. For n samples, must be repeated n times the
part of the sequence between parentheses (Berni PhD Thesys, modified).
Mercury has seven stable isotopes with the following approximate
abundances for the National Institute of Standards and Technology (NIST)
Standard Reference Material (SRM)-3133: 196
Hg = 0.155%, 198
Hg =
10.04%, 199
Hg = 16.94%, 200
Hg = 23.14%, 201
Hg = 13.17%, 202
Hg =
29.73%, 204
Hg = 6.83% (Blum and Bergquist 2007). These abundances are
referenced to the certified 205
Tl/203
Tl ratio of 2.38714 for the NIST SRM-
997.
The 202
Hg/198
Hg ratio will be used to describe mass-dependent stable
isotope fractionation (MDF) relative to the NIST SRM-3133. Hg isotopic
compositions relative to NIST SRM-3133 are reported as Eq. (6):
Chapter III: Methodology
57
1000*1)/(
)/(
3133
198
198
SRM
xxx
sample
xxx
HgHg
HgHgxxx
(6)
Mass-independent isotope fractionation (MIF) has been observed for the
odd isotopes of Hg and is reported as Δ199
Hg and Δ201
Hg. These values
represent the differences between the measured δ199
Hg and δ201
Hg values
and those predicted based on the measured δ202
Hg and the kinetic mass-
dependent fractionation law derived from transition state theory.
For variations of less than ~10‰, these values were approximated by Blum
and Bergquist (2007) to be:
)2520.0*( 202199199 HgHgHg
)5024.0*( 202200200 HgHgHg
)7520.0*( 202210201 HgHgHg
(7)
The magnitude of mass-dependent and mass independent fractionation
will be described primarily with δ202
Hg and Δ201
Hg.
In addition to variations in the magnitude of these isotope ratios, trends in
the ratios of the isotope ratios, in particular the Δ199
Hg/Δ201
Hg and
Δ201
Hg/δ202
Hg have been shown to be diagnostic of some chemical
transformation mechanisms (Bergquist and Blum 2007).
)493.1*( 202204204 HgHgHg
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
58
Chapter IV: Results
Chapter IV:
Results
59
4. Results
4.1 THg distribution in sediment and grain size
composition
The HgT contents measured in the core sediments show a wide range of
measures between 1.77 and 55.34 mgKg-1
(Tab. 9; Fig. 24). The lowest
values of THg are recorded in sediments collected in the northern Augusta
Bay (Fig. 9, station 7 replied in both sampling), ranging between 3.26 and
7.82 mgKg-1
(mean THg = 6.15±1.54 mgKg-1
) in May 2011 sampling and
replied in June 2012 (range: 4.12- 8.66 mgKg-1
; mean THg = 6.33±0.90
mgKg-1
) (Tab. 9; Fig. 24).
THg in sediments collected in the centre Augusta Bay (Station 11 and 12 of
2011 sampling and station 12 of 2012 sampling; Fig.24;) have intermediate
THg values, ranging between 13.79 and 17.98 mgKg-1
(mean THg =
16.23±1.46 mgKg-1
) and between 5.17 - 39.53 mgKg-1
(20.98±11.45 mgKg-
1) for core 11 and 12 of sampling 2011 respectively. THg content in core 12,
(sampling May 2012) range between 8.01 and 16.83 mgKg-1
(mean THg =
12.47±2.69 mgKg-1
).
Finally THg content in sediment collected in the southern Augusta Bay
(stations 16bis, sampling 2012; Fig. 24), has the highest THg values; range:
1.77 - 55.34 mgKg-1
(mean THg = 25.38±14.44 mgKg-1
). The HgT content
measured in core collected in station 16 in May 2011 (mean THg:
7.81±1.66) appears lower than those expected, with respect to the
geographical location (Fig. 24). The highest peaks of HgT contents were
measured in the southern collected cores, in core 12 (May 2011) at 13.5 cm
(39.53 mg Kg-1
) and in core 16bis (June 2012) at 15 cm (55.34 mg Kg-1
).
Sediments were texturally rather uniform at both sampling sites (Tab. 9;
Fig.25). Sediment layers consisted mostly of silt (~50%), and clay (~30%),
whereas the sandy fraction represents a small percentage (Tab. 9; Fig.25).
The only exception is represented by core 16 of May 2011 sampling, mainly
composed of sandy (~75%), thus justifying the minor THg content
measured in this core sediments (Tab. 9; Fig.25).
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
60
Figure 24: Profile of THg content along all the collected cores in 2011 and 2012, with
related sampling map.
0
3
6
9
12
15
18
21
24
27
30
33
36
0 10 20 30 40
Dep
th (c
m)
THg (mgKg-1) Sampling 2011
Core 7 Core 11 Core 12 Core 16
0
3
6
9
12
15
18
21
24
27
30
33
36
0 10 20 30 40 50 60
Dep
th (c
m)
THg (mgKg-1) Sampling 2012
Core 7
Core 12
Core 16 bis
Chapter IV: Results
61
Figure 25: Grain size composition in collected cores of sampling 2011 and 2012.
0
20
40
60
80
100
Sandy Silt Clay
Gra
in s
ize
fra
ctio
n (
%)
Core sediments 2011
St. 7 St.11 St. 12 St. 16
0
20
40
60
80
100
Sandy Silt Clay
Gra
in s
ize
fra
ctio
n (%
)
Core sediments 2012 St. 7 St. 12 St. 16 bis
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
62
Table 9: THg measures and grain size composition in all the collected sediments.
THg DHg are also reported for porewater sediment collected in June 2012.
Station Depth THg Mean
diameter Sorting Sandy Silt Clay
(cm) (mgKg-1) (μm) (μm) (%) (%) (%)
Sa
mp
lin
g M
ay
201
1
7
0- 3 7.48 9.11 5.02 8.14 59.91 31.95
3- 6 7.82 8.91 5.44 9.71 57.10 33.19
6 -9 7.53 8.72 5.23 8.69 58.03 33.28
9 -12 7.44 8.12 5.03 6.74 58.92 34.34
12 -15 5.96 8.72 5.34 8.61 57.37 34.02
15 -18 5.43 9.16 5.28 8.81 58.36 32.83
18 -21 3.26 8.25 5.43 8.70 56.95 34.35
21- 24 4.97 8.73 5.19 8.13 58.17 33.70
24 -26 5.41 9.04 5.22 8.77 57.90 33.33
11
0- 3 14.26 8.127 5.66 8.52 55.78 35.70
3-6 13.79 7.716 5.88 9.20 53.88 36.92
6-9 15.68 7.604 5.75 8.73 54.45 36.82
9-12 16.21 7.274 5.70 8.09 55.41 36.50
12-15 17.98 8.711 5.62 10.33 55.59 34.08
15-18 17.98 7.244 5.74 8.00 54.34 37.66
18-21 16.82 6.661 5.06 5.54 56.18 38.28
21-24 16.68 7.764 5.20 7.82 55.85 36.33
12
0-3 13.59 8.161 5.73 9.89 53.93 36.18
3-6 15.11 8.845 5.59 10.44 54.64 34.92
6-9 22.77 9.099 5.67 11.71 53.59 34.70
9-12 30.51 7.791 5.36 7.84 55.04 37.12
12-15 39.53 7.095 4.76 5.71 56.27 38.02
15-18 32.33 6.368 4.80 5.10 55.65 39.25
18-21 20.77 6.202 6.81 8.69 52.39 38.92
21- 24 9.02 8.875 6.00 11.96 52.35 35.69
24 -26 5.17 8.367 6.19 12.08 50.52 37.40
16
0-3 9.05 250.694 10.07 77.89 12.25 9.86
3-6 5.92 160.762 12.20 68.41 17.54 14.05
6-9 8.46 111.568 11.95 61.21 22.24 16.55
Sa
mp
lin
g J
un
e 2
01
2
7
0-1 8.34 4.24 6.64 3.27 52.65 44.08
1-2 6.07 4.80 6.41 4.00 54.62 41.38
2-3 6.79 4.45 6.61 3.46 53.86 42.68
3-4 6.04 4.61 6.63 3.74 54.39 41.87
4-5 6.42 4.43 7.26 4.67 52.62 42.71
5-6 6.31 4.51 6.20 3.30 53.98 42.72
6-7 6.16 4.52 6.54 2.01 55.66 42.33
7-8 6.55 5.64 6.84 6.35 55.03 38.62
8-9 5.66 7.27 7.26 10.27 55.20 34.53
Table 9 continued
Chapter IV: Results
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
63
9-10 5.88 5.80 7.10 6.34 55.91 37.75
10-11 6.09 5.16 7.04 4.96 55.61 39.43
11-12 6.15 3.89 7.35 3.27 52.18 44.55
12-13 5.99 4.56 6.42 2.74 54.67 42.59
13-14 8.66 4.73 7.37 6.13 52.17 41.70
14-16 6.84 5.00 6.36 2.47 56.54 40.99
16-18 6.67 4.67 6.59 3.96 53.36 42.68
18-20 7.23 5.93 6.16 5.48 56.25 38.27
20-22 6.72 5.85 6.66 6.77 54.29 38.94
22-24 7.15 6.28 6.21 6.68 55.25 38.07
24-26 5.74 5.14 7.02 5.13 60.07 34.80
26-28 4.12 4.77 7.10 4.18 54.69 41.13
28-30 5.08 6.54 7.19 10.19 52.45 37.36
30-32 5.59 5.84 6.96 6.53 55.37 38.10
32-34 6.32 5.46 6.70 6.12 55.17 38.71
34-36 6.35 4.56 7.11 5.16 52.47 42.37
36-37 5.64 4.77 7.04 3.57 55.27 41.16
12
0-2 15.96 7.37 5.98 8.79 53.71 37.50
2-4 13.36 8.07 5.74 8.98 55.51 35.51
4-6 13.61 4.95 7.93 8.22 49.84 41.94
6-8 12.16 4.91 7.07 7.01 50.74 42.25
8-10 16.83 6.53 5.96 6.53 54.61 38.86
10-12 12.80 6.57 6.23 7.81 53.40 38.79
12-14 14.24 6.13 6.50 7.38 53.04 39.58
14-16 9.33 6.61 6.18 7.69 52.89 39.42
16-18 12.63 4.57 6.92 5.30 51.60 43.10
18-20 11.83 4.72 6.54 3.68 53.76 42.56
20-22 8.01 8.10 5.91 9.11 55.68 35.21
22-24 9.12 5.40 7.69 9.14 49.81 41.05
16
0-2 17.11 8.10 7.09 12.72 51.20 36.08
2-4 16.90 9.05 7.20 15.84 49.52 34.64
4-6 18.99 5.45 7.87 10.57 47.13 42.30
6-8 22.69 5.66 7.75 11.23 47.22 41.55
8-10 24.94 6.63 6.56 9.55 50.03 40.42
10-12 28.13 5.03 6.98 7.96 47.68 44.36
12-14 38.41 5.82 7.40 9.80 48.90 41.30
14-16 55.34 5.69 6.51 8.77 48.69 42.54
16-18 46.67 6.42 6.84 9.97 49.17 40.86
18-20 39.96 7.10 6.75 9.91 52.04 38.05
20-22 35.71 6.05 6.82 7.93 51.97 40.10
22-24 31.12 5.93 7.81 11.05 48.41 40.54
24-26 18.51 6.95 7.28 11.21 51.43 37.36
26-28 28.69 7.40 7.11 12.16 50.70 37.14
Table 9 continued
64
28-30 22.08 6.81 7.01 10.32 51.71 37.97
30-32 7.28 7.14 7.90 13.55 49.75 36.70
32-34 2.57 10.61 6.72 15.84 52.26 31.90
34-36 1.77 9.03 6.62 11.86 54.65 33.49
4.1.1 Sequential extraction procedure
More than 80% of HgT in the sediments were extracted in step 4 (Tab. 10;
Fig. 26) thus documenting a highly stable phase of Hg, like Hg2Cl2
(calomel), or trapped in mineral lattice, or bound to humic substances
(Bloom et al., 2000). Around a 15% of Hg was extracted at the step 5,
possibly like insoluble Hg forms as cinnabar (HgS), m-HgS, HgSe and
HgAu. Only ~2% of HgT were released in the first three steps, as more
soluble, and bioavailable forms (Tab. 10; Fig. 26).
Table 10: THg extracted from each step applying the sequential extraction procedure on
sediments collected in May 2011.
Station Depth
(cm)
THg
(mgKg-1
)
STEP 1
(%)
STEP 2
(%)
STEP 3
(%)
STEP 4
(%)
STEP 5
(%)
7 6-9 7.53 0.206 0.059 1.472 84.832 13.431
21-24 4.97 0.048 0.040 0.276 95.267 4.370
11
0-3 14.26 0.039 0.026 0.300 82.106 17.528
3-6 13.79 0.031 0.031 1.001 89.804 9.1320
6-9 15.68 0.004 0.024 0.465 86.328 13.179
9-12 16.21 0.038 0.022 0.581 81.891 17.469
12-15 17.98 0.031 0.018 0.569 83.953 15.429
12-15 bis 17.98 0.021 0.119 0.813 93.190 5.857
15-18 17.98 0.026 0.080 0.541 86.240 13.112
18-21 16.82 0.039 0.024 0.641 80.037 19.259
21-24 16.68 0.038 0.018 0.545 82.895 16.504
24-26 16.66 0.042 0.017 0.559 87.156 12.225
12
0-3 13.59 0.042 0.025 1.037 81.293 17.603
0-3 bis 13.59 0.025 0.017 0.817 82.557 16.583
9-12 30.51 0.036 0.133 0.163 88.050 11.618
15-18 32.33 0.009 0.019 0.151 85.141 14.679
18-21 20.77 0.021 0.023 0.262 91.752 7.942
24-25 5.17 0.045 0.044 0.504 94.765 4.642
65
Figure 26: Hg (%) extracted from each step using the sequential extraction procedure.
4.2 Mercury in fishes
4.2.1 Total mercury concentrations in fishes
Hg analysis and bioaccumulation processes in fish compartment were
reported in attachment I (Bonsignore et. al., 2013).
Total mercury concentrations measured in tissues from pelagic, demersal
and benthic fishes, caught inside and outside of Augusta Bay, are
graphically summarised in Figures 27 a and b.
Mercury mean values calculated for each species, together with available
comparative data from the literature and HgT content measured in
anchovies from Marsala, are presented in Table 11.
Mercury concentrations ranged between 0.021 and 2.709 μg g-1 in muscles
(Fig. 27a) and between 0.029 and 9.720 μg g-1
in livers (Fig. 27b). The HgT
content in liver is from 1.5 to 6 times higher than that measured in muscles
from the same specimens (Table 11).
The highest HgT values were found in species caught inside the bay: 2
demersal specimens, a specimen of Diplodus vulgaris (HgT in liver = 4.979
μg g-1
) (extreme point in Fig. 27b) and a specimen of Serranus scriba (HgT
in muscle = 2.709 μg g-1
) (extreme point in Fig. 27a), a large pelagic
specimen of Sphyraena sphyraena (HgT = 9.720 and 2.269 μg g-1
in liver
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
66
and muscle, respectively) (Table 11) and a benthic specimen of Murena
helena (HgT = 2.638 μg g-1
in muscle) (Table 11).
However, these very high levels represent outliers of the whole dataset
(Figs. 27 a, b).
The highest non-outlier values refer once again to specimens caught inside
the bay and specifically to benthic species (Figs. 27 a, b). In particular,
Scorpanea scrofa and Scorpanea notata show the highest HgT mean
concentrations for both liver (1.638 and 2.339 μg g-1
, respectively) and
muscle (1.082 and 1.341 μg g-1
, respectively) (Table 11).
The lowest non-outlier ranges were found in pelagic specimens caught
outside the bay (0.021 - 0.167 μg g-1
for muscles and 0.029 - 0.5708 for
livers) (Figs. 27 a, b), and the HgT mean values measured in the different
studied species are substantially comparable (Table 11).
Finally, data for demersal species from the inner bay show the widest non-
outlier ranges (0.084 - 1.116 μg g-1
for muscles, 0.109 - 2.747 μg g-1
for
livers) and the most elevated number of outliers and extreme values (Figs.
27 a, b). In particular, the highest HgT mean values (2.165 μg g-1
in liver
and 2.581 μg g-1
in muscle) were measured in the Serranus scriba species
(Table 11).
Muscle
pelagic outside
pelagic inside
demersal outside
demersal inside
benthic outside
benthic inside
0.0
0.4
0.8
1.2
1.6
2.0
2.4
2.8
Hg
T c
on
cen
tra
tio
n
Sample
(µg g
-1)
Median
25%-75%
Non-Outlier Range
Outliers
Extremes
(a)
Chapter IV: Results
67
Liver
pelagic outside
pelagic inside
demersal outside
demersal inside
benthic outside
benthic inside
0.0
0.4
0.8
1.2
1.6
2.0
2.4
2.8
3.2
3.6
4.0
4.4
4.8
5.2
Hg
T c
on
cen
tra
tio
n
9,72
Sample(µ
g g
-1)
Median
25%-75%
Non-Outlier Range
Outliers
Extremes
(b)
Figure 27: Box-plots with HgT concentrations in the muscles (a) and livers (b) of pelagic,
demersal and benthic fishes.
Table 11: HgT means concentrations in muscle and liver of the analysed species and
comparison with data for other areas.
Species N°
Range of
total
length
(mm)
HgT
Muscle
(μg g-1)
S.D. References Site
HgT
Liver
(μg g-1)
S.D.
Engraulis
encrasicolus
40 109-138 0.052 0.019 This work Augusta 0.204 0.147
11 120-139 0.057 0.014 This work Marsala 0.119 0.038
0.040
Bilandžić et al.,
2011 Adriatic sea
9 121-147 0.070 0.090 Gibičar et al.,
2009 Adriatic sea*
0.030 0.030 Copat et al.,
2012 Sicily (Catania)
0.060 0.030 Copat et al.,
2012
Syracuse
(Sicily)
Table 11 continued
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
68
18
0.060
Pastor et al.,
1994
Mediterranean
sea (Spain)*
0.070
Martorell et al.,
2011
Mediterranean
sea (Spain)
4
0.055 0.003 Tuzen, 2009 Black Sea
(Turkey)*
Sardina
pilchardus
28 115-150 0.082 0.035 This work Augusta 0.196 0.157
10 168-178 0.090 0.040 Gibičar et al.,
2009 Adriatic sea*
0.080 0.030 Copat et al.,
2012 Catania (Sicily)
0.180
Buzina et al.,
1995 Adriatic sea*
0.198
Buzina et al.,
1995
Adriatic sea
(Kastela Bay)*
14
0.052
Wolfgang, 1983 Adriatic sea*
35 190-260 0.066
Wolfgang, 1983 Biscay Bay
5 157-165 0.050
Wolfgang, 1983 Mediterranean
sea
41
0.170
Wolfgang, 1983 Ligurian sea
28 120-150 0.030
Wolfgang, 1983 North Africa
(Ceuta)*
20 160-210 0.040
Wolfgang, 1983
Western
English
Channel
38
0.105
Pastor et al.,
1994
Mediterranean
sea (Spain)*
0.019
Martorell et al.,
2011
Mediterranean
sea (Spain)
7 188-200 0.033 0.016 Harakeh et al.,
1985 Lebanon
Boops boops
20 95-150 0.120 0.049 This work Augusta 0.236 0.191
11 158-198 0.196 0.204 Hornung et
al.,1980 Israel*
1
0.075
Pastor et al.,
1994
Mediterranean
sea (Spain)*
Table 11 continued
Chapter IV: Results
69
2 130-160 0.190
Stoeppler-
Nürnberg, 1979
Med. Sea
(Dubrovnik)
0.267
Buzina et al.,
1995 Adriatic sea
0.312
Buzina et al.,
1995
Adriatic sea
(Kastela Bay)*
16 139-171 0.036 0.025 Harakeh et al.,
1985 Lebanon
Trachurus
trachurus
6 56-222 0.131 0.147 This work Augusta 0.344 0.176
2 260-285 0.170
Stoeppler-
Nürnberg, 1979
North sea
(German Bight)
170 0.170
Mikac et al.,
1984
Adriatic sea
(Kastela Bay)*
37 130-236 0.122 0.101 Hornung et
al.,1980 Israel*
16 159-203 0.045 0.019 Harakeh et al.,
1985 Lebanon
0.053
Martorell et al.,
2011
Mediterranean
sea (Spain)
4
0.078 0.005 Tuzen, 2009 Black Sea
(Turkey)*
5
0.053 0.012 Keskin et al.,
2007
Marmara sea
(Turkey)*
Diplodus
annularis
74 109-179 0.557 0.303 This work Augusta 1.195 0.827
0.653
Buzina et al.,
1995 Adriatic sea*
0.628
Buzina et al.,
1995
Adriatic sea
(Kastela Bay)*
Diplodus
vulgaris
3 102-179 0.643 0.614 This work Augusta 2.035 2.554
5
0.378 0.017 Keskin et al.,
2007
Marmara sea
(Turkey)*
Sphyraena
sphyraena
1 1190 2.269
This work Augusta 9.727
14 219-295 0.167 0.068 Hornung et
al.,1980 Israel*
Caranx
rhonchus 1 264 1.701
This work Augusta
Table 11 continued
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
70
Pagellus
acarne
12 149-161 0.254 0.028 This work Augusta 0.618 0.178
3 135-141 0.112
Hornung et
al.,1980 Israel*
15 164-182 0.032 0.014 Harakeh et al.,
1985 Lebanon
Pagellus
bogaraveo 2 178-179 0.266 0.227 This work Augusta 1.230 0.700
Pagellus
erythrinus
8 154-205 0.407 0.100 This work Augusta 2.322 0.445
5 110 0.341 0.025 Papetti-Rossi,
2009
Tyrrhenian sea
(Lazio)
57 115-187 0.180 0.094 Hornung et
al.,1980 Israel*
9 89-173 0.240 0.190 Gibičar et al.,
2009 Adriatic sea*
28 140-152 0.042 0.023 Harakeh et al.,
1985 Lebanon
5
0.168
Pastor et al.,
1994
Mediterranean
sea (Spain)*
5
0.290 0.044 Keskin et al.,
2007
Marmara sea
(Turkey)*
Serranus
scriba
2 122-140 2.165 0.768 This work Augusta 2.581 0.592
3
1.030 0.459 Gibičar et al.,
2009
Tyrrhenian sea
(Tuscany)*
Mullus
barbatus
4 155-
202 0.815 0.777 This work Augusta 1.518 0.582
102-
230 0.116 0.056
Hornung et
al.,1980 Israel*
0.400 0.400 Storelli et al.,
2004 Ionian sea
0.490 0.500 Storelli et al.,
2004 Adriatic sea*
13 117-
180 0.700 0.730
Gibičar et al.,
2009 Adriatic sea*
0.370
Buzina et al.,
1995 Adriatic sea
0.318
Buzina et al., Adriatic sea
Table 11 continued
Chapter IV: Results
71
1995 (Kastela Bay)*
59
0.139
Pastor et al.,
1994
Mediterranean
sea (Spain)
0.010
Martorell et al.,
2011
Mediterranean
sea (Spain)
30 128-
166 0.054 0.025
Harakeh et al.,
1985 Lebanon
130-
200 0.233
Stoeppler-
Nürnberg, 1979
Mediterranean
Sea (Sardinia)
4
0.036 0.002 Tuzen, 2009 Black Sea
(Turkey)*
5
0.434 0.012 Keskin et al.,
2007
Marmara sea
(Turkey)*
Mullus
surmuletus
2 200-
209 0.662 0.089 This work Augusta 1.112
9 120-
160 0.086
Hornung et
al.,1980 Israel*
59
0.139
Pastor et al.,
1994
Mediterranean
sea (Spain)*
2 185-
203 0.250
Stoeppler-
Nürnberg, 1979
North sea
(German Bight)
37
0.060
Bilandžič et al,
2011
Adr.sea (Croatian
coast)
Scorpaena
scrofa
4 93-
112 1.082 0.285 This work Augusta 1.637 0.380
0.222
Buzina et al.,
1995 Adriatic sea
0.390
Buzina et al.,
1995
Adr. sea (Kastela
Bay)*
Scorpaena
notata
5 114-
133 1.340 0.380 This work Augusta 2.339 0.529
5
0.490 0.430 Gibičar et al.,
2009
Tyrrhenian sea
(Tuscany)*
Illex coindetii
6 33-92 0.078 0.039 This work Augusta
13 52-
224 0.100 0.100
Gibičar et al.,
2009 Adriatic sea*
Loligo forbesi 3 45- 0.147 0.024 This work Augusta 0.311 0.011
Table 11 continued
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
72
170
Sepia
officinalis 8
108-
148 0.766 0.288 This work Augusta
Octopus
vulgaris 1 123 0.443
This work Augusta
Murena
helena 1 800.5 2.638
This work Augusta 3.817
*Polluted site
4.2.2 Biological features
The fish caught from bottom-water sampling (inside the bay) consisted of 2
pelagic, 106 demersal and 16 benthic specimens, while specimens from
mid-water sampling (outside the bay), consisted of 103 pelagic, 3 demersal
and 1 benthic. Almost all the caught species, in particular, Engraulis
encrasicolus, Sardina pilchardus, Boops boops, Mullus barbatus and Illex
coindetii, are typical of the Mediterranean Sea and are commercially
relevant to Italian fishing (Irepa, 2010). Only one specimen was found to
belong to a so-called alien species, specifically Sphyraena sphyraena. This
is a typical species of the tropical seas, today present also in the
Mediterranean Sea (Streftaris and Zenetos, 2006). Assuming the length of
fishes as a reliable parameter for age estimates (Boening, 2000; Waldron
and Kerstan, 2001; Scudder et al., 2009; Panfili et al., 2010; Basilone et al.,
2011; Bacha et al., 2012) and, thus, reported HgT values vs. length to assess
biomagnification of that contaminant with time (Fig. 28a). Statistically
reliable and robust correlations were found between HgT mean values
measured in muscles for size classes and length in S. pilchardus (r2
= 0.75),
E. encrasicolus (r2
= 0.92), T. trachurus (r2
= 0.96), D. annularis (r2
= 0.98)
and P. erythrinus (r2
= 0.64) (Fig. 28a). Specifically, the calculated HgT
accumulation rates for S. pilchardus, E. encrasicolus, T. trachurus, P.
erythrinus and D. annularis are 0.011, 0.012, 0.022, 0.036 and 0.136 μg g-
1cm
-1, respectively, in good agreement with data reported by Hornung et al.
(1980) for P. erythrinus and T. trachurus species. In our dataset an evident
increasing trend was measured between HgT content and age in the two
most abundant species, E. encrasicolus and S. pilchardus (Fig. 28b) with
significant differences (p<0.005; ANOVA test) among age group, although,
the restricted range of available age classes needs a larger data collection.
Chapter IV: Results
73
Additionally, contamination effects show a south-north gradient evident
from HgT levels measured on the ubiquitous Pagellus spp. and D. annularis
specimens (Fig. 28c). In particular, the highest HgT mean concentrations
occur in fishes caught from southern Augusta Bay where bottom sediments
show the highest concentrations of mercury.
a)
b)
R² = 0.92
R² = 0.75
R² = 0.96
R² = 0.64
R² = 0.98
0
0.3
0.6
0.9
1.2
1.5
5 7 9 11 13 15 17 19 21 23
Hg
T (μ
g g
-1)
Size classes (cm)
Engraulis encrasicolus
Sardina pilchardus
Trachurus trachurus
Pagellus erythrinus
Diplodus annularis
0
0.02
0.04
0.06
0.08
0.1
0.12
0.14
0.16
0.18
0 1 2
Hg
T (μ
g g
-1)
Age (years)
Engraulis encrasicolus Sardina pilchardus
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
74
c)
Figure 28: Fish biological accumulation rate. a) Relationship between HgT concentrations
and total body length for Sardina pilchardus, Engraulis encrasicolus, T. trachurus, D.
annularis and P. erythrinus. Points represent the mean values for each size class; b)
Relationship between total mercury concentration (median value of HgT) in fish muscles
vs. age in E. encrasicolus and S. pilchardus. Black lines = confidence interval. c)
Differences in muscle (M) and liver (L) HgT contents in Pagellus spp. and Diplodus
annularis from the northern (C3) and the southern (C4) part of Augusta Bay.
4.3 Mercury (THg and DHg) in seawater
Mercury in waters was analyzed as total Hg (THg) and dissolved Hg (DHg),
the last one as filtered fraction, assuming that the difference between THg
and DHg approximate the content of Hg in suspended particulate matter
(PHg), as reported in other studies (e.g. Faganeli et al., 2003),
Reported THg and DHg values are calculated by difference with blank (10
ngL-) and measures lower than detection limit of the blank (calculated as
2.33σ) are reported as >L.d.. Figure 29 reported all the made analyses,
including values lower than the detection limits but close to the critical
limits (calculated as 4.65σ). The HgT concentrations in seawaters collected
during May 2011 (Tab. 12; Fig. 29) show a wide range of values, between
>L.d. and 129.27 ngL-1
(mean HgT: 20.24±26.18 ngL-1
) while HgD, ranges
Chapter IV: Results
75
between > L.d. and 21.3 ngL-1
with a mean value of 7.87±6.69 ngL-1
(excluding values < L.d.). An evident increasing trend of THg toward the
southern and more contaminated part of the Augusta Bay (where waste
spillage from chlor-alkali plant occurred) can be observed. In this zone very
high THg content (129 ngL-1
) has been found. On the vertical, surface
waters (20-30 m from the bottom, Fig. 29) show the lowest Hg content
(mean THg of 13.64±8.93 ngL-1
), with the minimum THg value of 0.57
ngL-1
, middle waters (10-20 m from the bottom; Fig. 29) have intermediate
concentration (mean THg of 24.36±30.9 ngL-1
) and deeper waters (10-20 m
from the bottom; Fig. 29) the highest content (mean THg of 51.87±55.09
ngL-1
). Basically that testifies testify a key role played by the sediments as
sources of mercury to the water column. THg in seawaters collected in June
2012 ranges between 0.45 and 14.85 ngL-1
(mean value: 7.43±5.19 ngL-1
)
and HgD mean of 2.26±2.24 ngL-1
with a maximum of 6.35 ngL-1
). Mean
THg results 4.48±4.65 ngL-1
in the northern Augusta Bay (station 7),
10.06±5.04 ngL-1
in the centre (station 12) and 11.31±9.07 ngL-1
in the
south area (station 16), in agreement with data reported the year before.
Concentrations of THg measured in seawater outside the Bay (Tab. 12;
station 18-21) range between 2.62 and 11.95 ng L-1
(mean: 6.46±2.95 ngL-
1), with no evident systematic change with depth. The highest HgT values
were measured closer to the bay (Fig. 9, station 19), with mean THg content
of 7.80±5.60 ngL-1
), while values measured in station far from the Augusta
Bay (Fig. 9; stations 19 and 20) are 5.15±0.83 and 6.44±2.37 respectively.
Unexpected values were recovered in open sea (Fig. 9; station 21) with
mean THg of 6.91±4.71 ngL-1
and DHg range between >L.d. and 5.55 ngL-1
(Tab. 12).
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
76
a) 10 m from the bottom
b) 10-20 m from the bottom c)20-30 m from the bottom
Fig. 29: Map of HgT content distribution in the Augusta Bay seawater created from May
2011 samples at 0-10 m from the bottom, 10-20 m from the bottom and from 20m to
surface water. Maps created by through3D block kriging and themed on the “natural
breaks” method.
Chapter IV: Results
Chapter IV: Results
77
Table 12: THg and DHg measurement in all the seawater samples collected inside (May
2011, June 2012) and outside the Augusta Bay (February 2012). Reported values are
calculated by difference with blank (10 ngL-); Lower measures are reported as >L.d.
Station Bathymetry Sampling depth THg DHg
(m) (m) (ngL-1
) (ngL-1
) S
am
pli
ng
Ma
y 2
011
1 14.0
11.2 17.77 < L.d.
6.2 9.17 6.70
1.4 9.17 3.20
2 16.0
10.7 17.67 < L.d.
6.7 14.87 < L.d.
2.2 0.57 0.30
3 11.0
8.4 29.97 0.30
4.6 11.97 1.30
2.3 0.57 0.30
4 8.0
3.1 6.27 3.20
0.1 0.57 < L.d.
5 20.0
15.9 17.67 0.30
9.2 7.07 < L.d.
1.0 9.17 < L.d.
6 17.0
13.5 20.57 < L.d.
6.7 3.37 < L.d.
2.0 6.27 < L.d.
8 22.0
19.2 14.87 7.30
9.5 15.87 14.30
1.0 4.27 3.20
9 22.0
18.2 23.47 3.30
10.2 14.87 < L.d.
1.4 14.87 < L.d.
10 27.0
23.4 19.27 11.80
13.5 3.37 3.20
1.6 17.67 3.20
11 24.0
21.9 127.07 < L.d.
11.5 129.27 19.80
1.2 26.27 7.50
13 32.0
29.3 3.37 < L.d.
16.9 12.67 8.90
2.4 17.67 3.20
14 21.0
16.3 2.27 < L.d.
9.4 18.67 8.90
1.0 11.97 6.00
15 26.0
20.6 57.77 20.30
11.2 20.57 1.10
2.4 23.37 21.30
16 22.0 16.5 11.97 14.60
Table 12 continued
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
78
11.3 23.47 14.60
0.5 3.37 20.30
17 16.0
12.7 34.87 3.20
7.3 31.97 14.60
1.6 22.57 6.00
Sa
mp
lin
g F
ebru
ary
20
12
18 17.0
8.0 11.05 5.55
2.0 4.55 < L.d.
19 42.0
27.0 6.10 3.55
16.0 4.80 < L.d.
2.0 2.45 1.25
20 85.0
81.3 9.15 3.55
45.3 5.42 < L.d.
6.1 4.75 0.55
21 679.0
679.0 11.95 4.15
375.0 6.15 5.55
20.0 2.62 <L.d.
Sa
mp
lin
g J
un
e 2
01
2
7 22.8
1.0 1.85 0.95
11.5 9.85 1.55
21.0 1.75 2.55
12 27.8
10.0 5.95 4.95
13.5 8.55 2.35
26.0 15.69 < L.d.
16 22.8
1.0 1.02 6.35
11.5 14.85 < L.d.
22.0 18.09 0.95
4.4 Atmospheric GEM distribution
Results related to GEM atmospheric assessment and wet and dry Hg
depositional flux were reported in (Tab. 12; Fig. 29) (Attachment II).
Briefly GEM distribution in the MBL results to be 1.5 ± 0.4 (range 0.9-
3.1) and 2.1 ± 0.98 (range 1.1-3.1) ng m−3
in the winter and summer,
respectively. The collected data are, however, somewhat higher than the
atmospheric background Hg level measured over the land at the
downtown urban site of Augusta (averaged 0.9±0.5 ng m-3
). Results on
wet deposition indicate Hg concentrations ranging from 21 to 32 ng l-1
(averaged: 26 ng L-1
).
Chapter IV: Results
79
Analysis carried out on dry deposition samples gave values ranging from
10.3 to 132.45 ng L-1
(averaged: 44 ng l-1
), indicating a variable amount
of suspended particulate matter into the atmosphere during the period of
surveying. The major input of Hg given by the particulate matter to the
chemistry of bulk deposition has been observed in samples collected on
April 2012, where we estimated a Hg contents of ~ 32 and 132.45 ng Hg
L-1
in wet and dry deposition, respectively.
A total mercury depositional flux (wet + dry) indicates values of Hg
bulk depositional flux ranging from 0.03 to 0.12 g m–2
day–1
(which
correspond to 11-44 g m–2
yr–1
) (Bagnato et al., 2013) .
4.5 Hg fluxes at the interfaces
4.5.1 Meteorological pattern of the area
A meteorological data set, including air temperature, wind intensity and
main direction, and precipitation amount, was developed using data from
the continuous acquisition by a weather station (DAVIS – Vantage Pro 2Wi-
Fi) installed on the roof of the Augusta port authorities office, sited in front
of the site location.
Meteorological condition during the benthic chamber experiments varies
during the 19-21 September 2011 (late summer-early autumn) and the 23-26
June 2012 sampling period (summer), with weather condition worst in the
first period (T: 23.5 °C; wind: 2.35 NW; precipitation: 0.15 mm) and
agreeable in summer (T: 26.4 °C; wind: 2.65 W-E; precipitation: 1.021
mm). Values do not represent monthly estimates, but are referred
exclusively to the sampling periods.
The differences in meteorological pattern of the two sampling period
(September 2011 and June 2012) allowed to consider seasonal variability in
the marine environment and to extend the measured fluxes to a yearly
estimation, although the last assessment may be overestimated because of
the absence of winter Hg flux.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
80
4.5.2 In situ benthic fluxes
Figure 30 shows DHg vs. time in water collected from the benthic chamber
during the two experiments in September 2011 and June 2012 (Fig. 9). The
in situ benthic flux in September 2011 in the northern part of the Bay
(station 7) shows a linear rise with time (r2= 0.53) ranging from 3.10 (t0) to
23.00 ngL-1
(t10) (Fig. 30a; Tab. 13).
On the contrary, trends of DHg content in the central Augusta Bay (station
11), (DHg range: 11.70 (t0) - 26.30 ngL-1
(t10)), are not linear with time.
Indeed the flux exhibits a rapid increase in the first hour (DHg= 35.40 ngL-
1), and a slow decrease with time (Fig. 30a; Tab. 13). A light, linear (r
2=
0.97) increment in DHg is observed in the southern station 16 (DHg range:
15.30 (t0) – 31.31 ngL-1
(t10).
In situ benthic flux in the northern Augusta Bay (St. 7) for June 2012
sampling, is comparable with the previous year, and increases linearly with
time, from <d.l (t0) to 6.34 ngL-1
at t4 and 29.99 ngL-1
at t10 (r2= 0.92; Fig.
30b; Tab. 13). DHg measured in station 12 (R2=0.77 ) quickly rises from
29.68 at t1 to 113.23 ngL-1
at t12.
Also in chamber 16bis the growth is linear (r2= 0.90) and range between
12.04 ngL-1
at t1 to 55.58 ngL-1
at t12 (Fig. 30b; Tab. 13). The calculated Hg
flux for the September 2011 sampling results 3, 8.1 and 8.4 μg m-2
day-1
for
station 7, 11 and 16 respectively, while for June 2012 sampling Hg flux
from each chamber results 8.1, 16.4 and 21.8 μg m-2
day-1
in station 7, 12
and 16 bis respectively.
The lower in situ benthic flux has been recorded in the north Augusta Bay
(st.7) while the highest in central and southern Augusta Bay.
Chapter IV: Results
81
a)
b)
Figure 30: Concentrations of HgD, vs. time during in situ flux chamber deployments in: a)
September 2011 and b) June 2012.
With the aim to estimate the total mercury flux from sediment to seawater
over the entire surface of the Augusta basin (~23.5 km2), we used the model
of territorial distribution proposed by Aurenhammer (1991) (Voronoi
Polygons method), calculating separate areas of the basin (in km2) which
were differently affected by Hg release from sediment.
Estimated fluxes from sediment to seawater for the whole Augusta Bay
result to be 0.22 kmol y-1
in 2011 (0.045 ty-1
) and 0.53 kmol y-1
in (0.11 ty-1
)
2012, more than one order of magnitude than those estimated by Sprovieri
et al. (2011) for the same study area. By mediating fluxes of the two years, a
total amount of 0.38 Kmol y-1
(0.08 ty-1
) has been obtained.
R² = 0.5305
R² = 0.9708
0
5
10
15
20
25
30
35
40
0 2 4 6 8 10 12
DH
g (
ng
L-1
)
Time (h)
September 2011 St. 7 St. 16 St. 11
R² = 0.9266
R² = 0.7719
R² = 0.9068
0
20
40
60
80
100
120
140
160
0 2 4 6 8 10 12
DH
g (
ngL
-1)
Time (h)
June 2012 St. 7 St. 12 St. 16 bis
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
82
Table 13: THg and DHg and standard deviation (3 replies) measured in water collected at
the interface sediment-water over time.
Sample Time
(h)
HgT
(ng L-1
)
D.S.
(n=3)
HgD
(ng L-1
)
S. D.
(n=3)
Flux
(μg m-2
day-1
)
Sa
mp
lin
g S
epte
mb
er
20
11
7
0 20.29 1.15 3.10
3.0 1 29.02 2.97 19.40
6 24.02 20.80 7.30
10 43.67 9.69 23.00 5.15
11
0 28.28 8.81 11.70 8.80
8.1 1 61.8 1.34 35.40 1.30
6 59.2 10.72 28.90 4.63
10 64.1 3.04 26.30
16
0 17.42 6.79 15.30 2.50
8.4 1 21.52 8.27 16.20
6 25.52 5.66 22.22
10 36.52 31.32
Sa
mp
lin
g J
un
e 2
01
2
7
0 15.96 19.07 0 10.03
8.1
1 20.80 4.39 0 16.19
4 39.34 5.46 6.34 2.05
6 43.24 4.30 21.16 25.7
12 49.86 16.26 29.99 54.65
12
0 70.89 24.87 29.68 5.66
16.4
1 95.50 0.05 69.73 4.10
4 137.06 1.84 86.70 60.10
6 145.66 14.00 95.43 44.53
12 152.78 14.40 113.23 29.69
16
0 19.16 0
21.8
1 38.96 58.97 12.04 37.73
4 51.86 40.73 34.1 18.36
6 70.41 14.50 38.68 35.92
12 84.86 55.58 28.99
Chapter IV: Results
83
Table 14: Calculated fluxes in Augusta Bay compared with those reported for other
polluted site.
Period Site Φ
μg m-2
day-1
Ref.
September 2011
7 3.0
This work
11 8.1
16 8.4
June 2012
7 8.1
12 16.4
16bis 21.8
Gulf of Trieste 5-6 Covelli et al., 1999
Bellingham Bay 10 Bothner et la., 1980
Arcachon Bay 5 Bouchet et al., 2011
Grado Lagoon 37-77 Covelli et al., 2008
4.5.3 Air-sea GEM Flux
Flux assessment at the sea-air interface was reported by Bagnato et al., 2013
(Attachment II). The estimated mercury evasion fluxes range from 3.6
(unpolluted site) to 36 (most polluted sites) ng m-2
h-1
which correspond to
an annually mercury evasion rate of ~ 31.5 and 315 mg m-2
yr-1
, respectively
(Bagnato et al., 2013).
The sea-air evasion flux of Hg from the basin is not uniformly distributed
but varies spatially, while any particular trend across the two seasons
(November 2011-June 2012) has been observed.
The higher Hg evasion flux was estimated in the southern part of the basin,
which corresponds to the most contaminated area of the basin in term of Hg
contained in the bottom sediments.
The lowest Hg evasion flux has been estimated in the northern sector of the
basin, close to the port, where the bottom sediments contain the lower Hg
concentrations found.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
84
4.6 REEs distribution in seawater
REE concentration in seawater collected in Augusta Bay ranges between
46.7 and 255.7 ngL-1
.
The highest REE contents were recorded for light REE (LREE) from La to
Nd, with an unusual value for Ce. These values are maximum in the middle
and deeper water column (between 11 and 25 m), and decrease in surface
waters. A similar distribution has already been observed in the
Mediterranean coastal areas and attributed to the leaching of REE from
atmospheric particulate once in contact with the blade surface water (Censi
et al., 2007; 2010).
As well as REE concentration differs in bottom water, also the PAAS-
normalized distribution has the same trend (Fig. 31). Surface samples (Fig.
31a) have a generically increasing trend along the series, with a light
enrichment of intermediate REE (MREE). In bottom water REE
concentrations and their trend seem to be similar in station 12 and 16, while
differ in station 7, in which REE concentration is lower. Gd anomaly
(Gd/Gd*>1) was recorded in all the station for the entire water column, with
the highest anomaly recorded in middle and deeper water.
Contrary to those expected in a marine water, in which Ce/Ce* values are
generally <1 (de Baar et al., 1983), no Ce negative anomaly has been
recorded.
Positive relationship was found by plotting Gd vs. Ce anomalies in the
analysed samples (r2=0.88) (Fig. 32), testified common anthropic process
and consequent liquid waste discharge into the water column of the Augusta
Bay. The Y/Ho ratio is commonly used as proxy of terrestrial contribution
to sediments, because Y behaves in accord with Ho and no fractionation
between the two elements occurs terrestrially, while Y fractionates from Ho
in marine reaction systems. In the considered dataset, the Y/Ho molar ratio
ranges between 55.2 and 95.0, with the higher values in surface water and
lower in depth. Y/Ho value in station 7 are the closest to chondrites
signature (molar ratio: 52), while ratio in station 12 is higher, especially in
surface waters. This clearly indicates the leaching of element by lithogenic
particulates, that justifies the crustal signature of station 7. This indication is
less strong in station 16 and absent in station 12. In this station the Y/ho
ratio suggests a preferential Ho removal, respect to Y, a quite common
Chapter IV: Results
85
process in marine environment (Bau & Dulski, 1996; Bau et al., 1996; Alibo
& Nozaki, 1999b).
Figure 31: Shale-normalised (PAAS) REE patterns of the analysed seawater samples at
different depths. a) surface water; b) middle water; c) bottom water.
1.00E-08
1.00E-07
1.00E-06
1.00E-05
La Ce Pr Nd Sm Eu Gd Tb Dy Ho Er Tm Yb Lu
[RE
E]
(n-P
AA
S)
a St. 7 St. 12 St. 16
1.00E-08
1.00E-07
1.00E-06
1.00E-05
La Ce Pr Nd Sm Eu Gd Tb Dy Ho Er Tm Yb Lu
[RE
E]
(n-P
AA
S)
b
1.00E-08
1.00E-07
1.00E-06
1.00E-05
La Ce Pr Nd Sm Eu Gd Tb Dy Ho Er Tm Yb Lu
[RE
E]
(n-P
AA
S)
c
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
86
Figure 32: Relationship between Ce/Ce* and Gd/Gd* anomalies.
Chapter IV: Results
87
4.7 Hg isotopes analyses
4.7.1 Hg isotopic composition of fishes
Mass dependent fractioning in fishes, expressed as δ202
Hg values,
exhibits a wide range of measures (-1.52 to 0.37 ‰) with a negative
mean δ202
Hg value of -0.36 ‰ (Tab.15).
Negative MDF fractioning are recorded especially in demersal and
benthic species (Pagellus spp, and Mullus barbatus), with a unique
exception for a specimen of Sardina pilchardus, considered as outlier.
Position of fishes in the water column appears to be the major driver of
the range in δ202
Hg. Mass independent fractioning in fishes exhibits
values positive or close to zero (range: -0.07 to 1.71 ‰ for Δ199
Hg and
0.02 to 1.48 ‰ for Δ201
Hg) (Tab. 15). Because the fish tissues display
MIF, it is instructive to also plot the data as Δ199
Hg vs. Δ201
Hg. The
Δ199
Hg vs Δ201
Hg ratio is diagnostic of the reduction of Hg (II) vs.
MeHg. A good correlation was obtained by plotting Δ199
vs. Δ201
(r2=0.84), and the slope ~1, indicating Hg(II) photoreduction (Fig. 33 a).
Difference among species emerged when the fish dataset was divided
according to the habitat depth that fish occupy (Fig. 33 b).
Contrary to benthic fishes, where Δ199
Hg and Δ201
Hg are close to zero
(0.09 ‰ and 0.11 ‰ respectively), MIF increases in demersal (0.23 ‰
for Δ199
Hg and 0.29 ‰ for Δ201
Hg) and displays the maximum
fractionation in pelagic fishes (1.2 ‰ for both Δ199
Hg and Δ201
Hg) (Fig.
33 b).
MeHg displays more MIF compared to Hg (II) and the ratio Δ201
Hg vs
δ202
Hg is predictive of photos or demethylation reduction. The calculated
ratio Δ201
Hg/δ202
Hg for fishes results to be 1.7.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
88
Fig. 33: Linear correlations between Δ199
Hg and Δ201
Hg (‰) for the entire fish dataset
(a) and divided into pelagic, demersal and benthic fishes (b).
y = 1.0033x - 0.0281
R² = 0.84
-0.2
0
0.2
0.4
0.6
0.8
1
1.2
1.4
1.6
1.8
-0.3 0.2 0.7 1.2 1.7
Δ199 H
g
Δ201 Hg
a)
-0.2
0
0.2
0.4
0.6
0.8
1
1.2
1.4
1.6
1.8
-0.3 0.2 0.7 1.2 1.7
Δ1
99 H
g
Δ201 Hg
b) Pelagic
Demersal
Benthic
Chapter IV: Results
89
4.7.2 Sediment Hg isotopic fractionation
Mercury isotopes were first measured in marine sediments by Gehrke et al.
(2009) in a study of mid- Pleistocene (955 kyr) Mediterranean sediment
cores and reported the Hg isotopic composition of sediments of δ202
Hg = -
0.91‰ and Δ201
Hg = 0.04‰, providing the first estimate of a
preanthropogenic marine Hg isotope composition. The Augusta Bay
sediment cores display negative δ202
Hg values, ranging between -0.91 and -
0.03‰ (Tab. 15). As shown in Fig. 34a, MDF intensity varies in core
collected in the tree station points; in core 7 (northern Augusta bay) the
mean δ202
Hg results -0.26±0.12 ‰, while decreases in core 12 (middle
station) (mean δ202
Hg: -0.36±0.17 ‰) and reaches the minimum value in
core 16, collected in the southern part of the Bay (δ202
Hg: -0.52±0.25 ‰).
On the other hand, MIF fractionation displays the opposite trend (Fig. 34 b),
with higher values in the southern part (station 16) (Δ199
Hg: -0.004±0.02‰),
intermediate in the middle (Δ199
Hg: -0.02±0.02‰) and the lowest values in
the northern station 7 (Δ199
Hg: -0.04±0.02‰). The Δ201
Hg mean value for
all the analyzed sediments results to be 0.02±0.03 ‰. By plotting δ202
Hg vs
Δ201
Hg for the entire sediment core dataset, on obtain a peculiar trend (Fig.
35), corresponding, according to the scheme proposed by Blum (2011) and
Rodriguez-Gonzales et al., (2009), to biological mediated MeHg production.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
90
MDF in sediments
Mean
Mean±SE
Mean±SD North Middle South
-0.8
-0.7
-0.6
-0.5
-0.4
-0.3
-0.2
-0.1δ
202H
g
b) MIF in sediments
Mean
Mean±SE
Mean±SD North Middle South
-0.07
-0.06
-0.05
-0.04
-0.03
-0.02
-0.01
0.00
0.01
0.02
0.03
Δ199H
g
Figure 34: Box plots of Hg isotopes fractionation in sediment core collected in north,
middle and south Augusta Bay.
a) MDF fractionation, expressed as δ202
Hg; b) MIF fractionation, expressed as Δ199
Hg.
Chapter IV: Results
91
Figure 35: Relationship between Δ
201Hg and δ
202Hg in sediments.
4.7.3 Hair Hg isotopic composition
MDF values in hair samples range between 1.15 and 3.36 ‰ (mean δ202
Hg=
1.17±0.48‰) (Tab.15) with no evident shift with age. Also MIF values are
positive (range: 0.73-1.51‰, mean 1.18±0.17‰ for Δ199
Hg; range: 0.80 -
1.33; mean 0.97±0.16‰ for Δ201
Hg). The Δ201
Hg trend shows a slight
enrichment with increasing age, with Δ201
Hg of 0.83 ‰ at 20 year (n=1),
0.88 ‰ at 30 years (n=2), 0.93 ‰ at 35 year (n=4) and 1.01 at 40 years.
A positive correlation has been obtained by plotting Δ201
Hg vs Δ199
Hg
(r2=0.87), with a slope of 1.2 compatible, according to the scheme proposed
by Bergquist and Blum (2007), with photochemical reaction (Fig. 36).
-2.0
-1.0
0.0
1.0
2.0
-1.0 0.0 1.0
Δ201H
g (
%0)
δ202 Hg (%0)
Sediment
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
92
Figure 36: Relationship between Δ
201Hg vs Δ
199 Hg (r
2= 0.87; slope: 1.2) in human hair
living in Augusta Bay.
Table 15: Details on Analysed fish and sediment samples, THg content and Hg isotopic
fractioning (‰)
Range
δ202
Hg
Mean
δ202
Hg
Range
Δ199
Hg
Mean
Δ199
Hg
Range
Δ201
Hg
Mean
Δ201
Hg
FISH
(Pelagic)
-1.52 to
0.07 -0.36 0.71 to1.71 1.21 0.66 to 1.48 1.21
FISH
(Demersal)
-1.27 to
0.1 -0.51
-0.02 to
0.58 0.23 0.05 to 0.42 0.29
FISH
(Benthic)
-0.75 to
0.37 -0.13
-0.02 to
0.55 0.09 0.05 to 0.64 0.11
ALL
FISHES
-1.52 to
0.37 -0.39
-0.07 to
1.71 0.65 0.02 to 1.48 0.67
SEDIMENT -0.91 to -
0.03 -0.38
-0.11 to -
0.02 -0.02
-0.37 to
0.04 -0.032
HAIR 1.15 to
3.36 1.71
0.73 to
1.51 1.18 0.80 to 1.33 0.97
y = 1.2064x
R² = 0.87
-1.0
-0.5
0.0
0.5
1.0
1.5
2.0
-1.5 -1.0 -0.5 0.0 0.5 1.0 1.5
Δ199 H
g (
%0)
Δ201 Hg (%0)
MIF in hair
Chapter IV: Results
93
Chapter V: Discussions
Chapter V:
Discussions
94
5. Discussion
5.1 Pattern of Hg distribution in the Augusta Bay
In the last decade, several studies (ICRAM, 2005; Ausili et al., 2008; Di
Leonardo et al., 2007, 2008; ENVIRON International Team, 2008; Ficco et
al., 2009; Sprovieri et al., 2011) have provided detailed information on the
pollution levels and risks for human health of resident populations of
Augusta Bay, an area, located in the eastern Sicilian coast, strongly
contaminated due to the uncontrolled chemical discharges from the most
complex petrochemical district in Europe.
Here the extremely high concentrations of Hg measured in the Augusta Bay
sediments, more than one order of magnitude higher than background
literature (Di Leonardo et al., 2006, Ogrinc et al., 2007) confirm the warning
previously reported and call for more detailed study on the biogeochemical
dynamics of Hg in the study area, in order to better understand the effect of
this pollution on the surrounding environmental compartments.
5.1.1 Spatial Hg distribution of THg content in sediment
A sud-north gradient of Hg exposure is testified by the extremely high Hg
content measured in core sediments collected in the southern area, closest to
the “Vallone della neve” where discharge from the alkali-plant occurred in
the recent past. As known (Devai et al., 2005; Vane et al., 2009; Feyte et al.,
2010), anthropogenic mercury has often been reported to be associated with
both organic matter and fine-grained particles of bottom sediments, because
of the tendency of fine-grained particles to bind large amounts of trace
metals due to their chemical composition and great surface to volume ratio
(Förstner and Wittman, 1979; Ravichandran, 2004).
Sediments were texturally rather uniform at both sampling sites and is
primary composed of silt and clay, whereas the sandy fraction represents a
small percentage. The only exception is represented by core 16 of May 2011
sampling, mainly composed of sandy, thus justifying the minor THg content
measured in this core, lower than expected respect of the geographical
position.
Hg speciation in contaminated sediments is crucial in order to understand
Hg mobility and bioavailability (Davis, et. al, 1997). For this reason, the
sequential selective extractions procedure (SEP), validated from Bloom, et
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
95
al, 2000, has been applied, in order to differentiates Hg into behavioral
classes including water soluble, ‘stomach acid’ soluble, organochelated,
elemental, and mercuric sulfide.
Although this method do not provide species-specific information, it does
offer the opportunity to investigate on the biogeochemically relevant
fractionation of Hg in sediments. Sequential extraction procedure
demonstrate the most part of HgT in sediments (~80% of the total) consists
of highly stable phase of Hg, like Hg2Cl2 (calomel), or trapped in mineral
lattice, or bound to humic substances, while ~15% of Hg were like insoluble
Hg forms (as cinnabar (HgS), m-HgS, HgSe and HgAu), and only ~2% of
HgT is in more soluble, and bioavailable forms. This is due to the tendency
of Hg2+
to bind with the sulfur-groups is notable, especially where the
concentration of HS- and S
2- is high, as in anoxic waters and sediments,
where mercury sulfide, solid and slightly soluble in water, tends to
accumulate into the sediments (Langer et al., 2001; Gilmour and Henry,
1991).
Some anaerobic microorganisms can manage these stable Hg forms (HgCl2,
Hg(SH)2, HgS and Hg0 trapped in minerals structures) as substrate for their
metabolism, this making Hg more easily bioavailable for the environment
(Benoit et al., 1999 a, b, 2001 a, b).
Further investigation, regarding the analysis of other useful parameters such
as dissolved Fe and Mn and H2S and MMHg, will permit to have a
complete view of the biogeochemistry.
5.1.2 Spatial Hg distribution of THg content in seawater
A north-south gradient of Hg distribution was confirmed by THg and DHg
analysis in seawater. The highest values were recorded in the southern and
more contaminated part of the Augusta Bay, in good agreement with Hg
distribution in sediments.
A clear gradient was also observer along the water column, with higher
values in deeper water and lower in surface.
This vertical trend is clearly owing to the major remoteness to polluted
sediments and to volatilization processes, occurring in surface water when
water column is undersaturated in the most volatile and long-lived Hg form,
Hg0
(Fitzgerald and Clarkson, 1991). Moreover, the THg and DHg
measured in seawater collected outside are anomalous compared to
Mediterranean sea (0.2 e 0.4 ngL-1
) (Kotnik et al, 2007; Horvat et al, 2003;
Chapter V: Discussions
96
Rajar et al., 2007; Cossa et al., 1997), and exhibit the maximum values
close to the two narrow inlets: the Scirocco and the Levante ones.
5.1.3 Mercury bioaccumulation pattern in fishes
The HgT content measured in fish tissues from Augusta Bay show an
increasing trend with habitat depth, specifically, with highest values
measured in benthic species, intermediate in demersal fishes and lower in
pelagic organisms. In particular, the highest HgT mean concentrations occur
in fishes caught in the southern Augusta Bay where bottom sediments show
the highest concentrations of mercury.
Additionally, contamination effects show a south-north gradient evident
from HgT levels measured on the ubiquitous Pagellus spp. and D. annularis
specimens. In particular, the highest HgT mean concentrations occur in
fishes caught from southern Augusta Bay where bottom sediments show the
highest concentrations of mercury.
The Hg accumulation in marine fish primarily depends on some important
biokinetic parameters: assimilation from the ingested prey, uptake constants
from the aqueous phase, de-toxification rate (Wang, 2012; Wang et al.,
1997, 1998; Wang and Fisher, 1999; Dang and Wang, 2011) and
environmental features (e.g., Hg concentration and speciation in seawater,
dietary sources, etc.) (Wang and Wong, 2003).
However, physiological and geochemical species-specific influences on Hg
bioaccumulation are still not fully understood (Baines et al., 2002; Xu and
Wang, 2002; Wang and Wong, 2003; Dang and Wang, 2012). Several
studies have demonstrated that Hg concentrations in the muscles of marine
organisms proportionally increase with size and age (Lange et al., 1994;
Burger et al., 2001; Green and Knutzen, 2003; Simonin et al., 2008).
Moreover, Hg de-toxification rates appear negatively correlated with the
fish size (Trudel and Rasmussen, 1997), supporting a potential correlation
between Hg levels and size/age in the organisms.
However, detailed investigations on different groups of species and on a
wide range of HgT concentrations are lacking and, when available,
sometimes controversial (Stafford and Haines, 2001), especially for fishes
with low mercury levels (average below 0.2 ppm) (Park and Curtis, 1997;
Burger and Gochfeld, 2011). Strong correlations between size and Hg levels
in fish are reported for Swordfish (Xiphias gladius) and Bluefin Tuna
(Thunnus thynnus) from the Mediterranean Sea (Storelli and Marcotrigiano,
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
97
2001), for several pelagic fish species from the Adriatic Sea (Storelli, 2008)
and for S. pilchardus specimens from Tunisia (Joiris et al., 1999).
Furthermore, Burger et al. (2007) found a positive correlation between size
and Hg levels for 11 of 14 species of marine fishes collected in the western
Aleutians (Bering Sea/North Pacific) and Luten et al. (1987) found the same
positive correlation in Atlantic Cod. Moreover, Leonzio et al. (1981) report
positive correlations between Hg content and weight in M. barbatus and a
slight Hg increasing trend with size in E. encrasicolus from the northern
Tyrrhenian Sea.
Finally, Gewurtz et al. (2011) show strong correlation between HgT
concentration and length in most freshwater fishes from the Canadian Great
Lakes and Ontario (Canada).
Here, the high number of specimens available from pelagic, benthic and
demersal fish species associated with a wide range of length/age and HgT
variability detected in tissues offer a challenging opportunity to explore in
more depth the actual bioaccumulation processes of Hg in the studied
organisms. In particular, we assumed the length of fishes as a reliable
parameter for age estimates (Boening, 2000; Waldron and Kerstan, 2001;
Scudder et al., 2009; Panfili et al., 2010; Basilone et al., 2011; Bacha et al.,
2012) and, thus, reported HgT values vs. length to assess biomagnification
of that contaminant with time.
Statistically reliable and robust correlations were found between HgT mean
values measured in muscles for size classes and length in S. pilchardus, E.
encrasicolus, Trachurus trachurus, Diplodus annularis and P. erythrinus.
Specifically, the calculated HgT accumulation rates for S. pilchardus, E.
encrasicolus, T. trachurus, P. erythrinus and D. annularis are 0.011, 0.012,
0.022, 0.036 and 0.136 μg g-1
cm-1
, respectively, in good agreement with
data reported by Hornung et al. (1980) for P. erythrinus and T. trachurus
species.
This definitively supports a significant linear HgT-length relationship for
the studied fish species and a species-specific accumulation effect on the
studied marine organisms. In our dataset an evident increasing trend was
measured between HgT content and age in the two most abundant species,
E. encrasicolus and S. pilchardus with significant differences (p < 0.005;
ANOVA test) among age group, although, the restricted range of available
age classes needs a larger data collection.
Muscles are the most commonly analysed tissues to monitor Hg levels in
fishes because they represent the edible part of the organism associated with
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
98
human health risk implications (Henry et al., 2004). Indeed, Hg accumulates
over time more readily in liver than in muscle, but muscle appears to retain
Hg for a much longer period (Boudou and Ribeyre, 1995). Thus, liver may
provide information only on short-term exposure to Hg pollution or may
bioaccumulate only when an organism is exposed to constant or increasing
levels of dietary mercury (Atwell et al., 1998). This is clearly reflected in
the studied dataset, where HgT concentration measured in liver is up to two
orders of magnitude higher than in muscles.
A direct comparison of HgT content in benthic species from Augusta Bay
and other Mediterranean areas affected by comparable Hg discharges by
chlor-alkali plant and sewage sludge disposal, specifically Tuscany and
Israel (Hornung et al., 1980; Gibičar et al., 2009), show 2–7 times higher
values, thus, underlying the combined effects of high pollution levels and
specific biogeochemical pathways driving mercury bioavailability in the
studied system.
The mean HgT concentrations measured in pelagic species caught outside
the bay, are similar to those reported for other sites affected by Hg
pollution: the Adriatic Sea (Wolfgang, 1983; Storelli and Marcotrigiano,
2001; Storelli et al., 2002, 2004, 2007, 2010; Gibičar et al., 2009), Turkish
areas (Tuzen, 2009), Spanish coastal areas (Pastor et al., 1994) and Israel
area (Hornung et al., 1980). Particularly the HgT mean concentrations
measured in the livers of E. encrasicolus specimens from Augusta Bay are
about twice as high (p = 0.044) as those measured in fishes from the
unpolluted area of Marsala, suggesting a direct, short-term effect of the bay
pollution on the pelagic fishes.
5.2 Toxicological effect on fish compartment
Although estimation of the target hazard quotient (THQ) and weekly intake
(EWI), do not provide a quantitative estimate on the dangerous health
effects on exposed populations, many authors implies these methodologies
as a preliminary information on the health risk level resulting from pollutant
exposure (Storelli et al., 2005, 2010; Storelli, 2008; Martorell et al., 2011;
Domingo et al., 2012) .
Target hazard quotient (THQ) and estimated weekly intake (EWI) were
calculated for muscles of fishes caught inside and outside the bay. The
Chapter V: Discussions
99
Target hazard quotient was calculated according to the US EPA (1989)
method and it is described by the following equation:
310
TAWABRFD
CFIREDEFTHQ
where EF is exposure frequency (365 days/year); ED is the exposure
duration (70 years), equivalent to the average lifetime; FIR is the food
ingestion rate (36 g/person/day) (FAO, 2005); C is the metal concentration
in seafood (μg g-1
); RFD is the USEPA’s reference dose (0.1 µg Hg kg bw-1
d-1
) (http://cfpub.epa.gov) or acceptable daily intake determined by WHO,
2003 (0.23 µg Hg kg bw-1
d-1
) (http://apps.who.int); WAB is the average
body weight (60 kg), and TA is the average exposure time for no
carcinogens (365 days/year x ED). The THQ was calculated for all the
studied species in the Augusta Bay using the US-EPA’s reference dose
(THQa) and the acceptable daily intake determined by the WHO (THQb).
These two values are relative to MeHg, nevertheless, we used them basing
on the assumption that the majority, if not all, of the Hg bioaccumulated
through the food chain is MeHg (Winfrey and Rudd, 1990; Mason and
Fitzgerald, 1990, 1991; Gilmour and Henry, 1991; Horvat et al., 1999;
Carbonell et al., 2009).
The estimated weekly intake (EWI) was calculated by multiplying the HgT
concentration (C) times by the weekly dietary intakes (FIR x 7) and
reporting to the average body weight (WAB). Finally, mean THQ and EWI
values were calculated for each studied species. Also, considering that
fishing activity within the bay has been interdicted since 2007 (Order N°
73/07), data relative to fishes from inside and outside the bay were
processed separately. Mean THQ and EWI values calculated for each caught
species, inside and outside the bay, are reported in Bonsignore et al., 2013
(attachment I). In particular, for no carcinogenic effects, an HQ exceeding
1.0 indicates a potential health risk (US EPA, 1989). In our dataset, either
using USEPA’s reference dose (TQa) that WHO acceptable daily intake
(TQb), species inside the Bay exceeded the value 1 in all cases, while fishes
outside the Bay only in demersal and benthic fishes (P. bogaraveo, P.
erythrinus, M. barbatus).
International Agencies indicate a provisional tolerable weekly intake
(PTWI) of Hg, representing safe values for human population over lifetime,
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
100
ranging from 0.7 μg kg-1
body weight (b.w.) (US-EPA, 2004) to 1.6 μg kg-1
b.w. (FAO/WHO, 2006).
The calculated EWI index exceed the PTWI (US-EPA, 2004; FAO/WHO)
in almost all the species collected inside the Bay and in demersal and
benthic fishes from outside, thus advising that consumption of fish from
inside the Augusta Bay represents a serious risk for human health of the
local populations, while suggest caution in consuming demersal and benthic
fishes from outside the Augusta Bay definitively demanding for appropriate
social actions.
5.3 Sediments as primary source of Hg
The high THg concentrations measured in benthic species, together with the
anomalous THg content measured in pelagic species living outside the Bay,
suggested active Hg release mechanism from polluted sediments to the
water column, with consequent effects of bioaccumulation in the trophic
web.
Also the THg and DHg distribution in water column confirm the hypothesis
of sediments as primary sources for the surrounding environment. Indeed
the reported high resolution maps of THg distribution in column water
demonstrate a direct relationship between THg and DHg contents in
seawater and THg content in sediments thus suggesting the key role played
by the highly polluted sediments as sources of Hg to the investigated marine
environment, as speculated by Sprovieri et al., 2011.
The applied benthic chamber experiment offered the unique opportunity to
measure the authentic release of Hg from Augusta Bay sediment to seawater
over time. The calculated fluxes at the interface between sediment and
seawater in Augusta Bay are higher than those reported for the Trieste Gulf
(Covelli et al., 1999), the Bellingham Bay (Bothner et al., 1980), the
Arcachon Bay (Bouchet et al., 2011).
Calculated fluxes are linearly correlated (R2=0.68) with mean THg content
in the first 5 cm of sediments (Fig. 37), thus suggesting an increasing of Hg
content in sediments determine a major release flow. The north-south
gradient before reported for fishes, seawater and sediments, is also reflected
by the in situ benthic fluxes. Indeed, the lowest diffusive fluxes were
recorded in the northern Augusta bay (St.7), intermediate fluxes in the
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
101
middle zone (St. 11 and 12) and the highest in the southern part (St. 16bis),
as measured for sediment (Fig. 37).
The only exception is represented by station 16 (sampling September 2011),
reflecting the lower THg content in sediment, due to the sandy nature of this
core.
Figure 37: Relationship THg content in sediments, calculated for the first 5 cm, vs. benthic
fluxes in the same station point for the entire dataset( September 2011 sampling, yellow
bubble and June 2012 sampling, green bubble).
With the exception of station 16 (for the above evidence), higher Hg fluxes
were connected to the higher temperature in warmer periods (June 2012) at
both sites suggesting that temporal variations of these fluxes are temperature
dependent (Fig. 37). Indeed during warmer period, the highest microbial
activity, and bottom-water O2 content enhance the importance of metal
reductions in sedimentary biogeochemical processes (Covelli et al., 2008).
Moreover higher temperatures enhancing microbial activity that, in turn,
stimulate Hg methylation (Ullrich et al., 2001).
Moreover the here used dynamic flux chamber technique, coupled with a
real-time atomic adsorption spectrometer, has represent an important step
aimed to refine the estimation method to assess Hg fluxes from
environmental surfaces (Wang et al., 2006).
Data of fluxes at the air-sea interface, result to be higher than Hg flux from
background uncontaminated soils (Ericksen et al., 2006), while are
comparable to those reported from volcanic/geothermal areas (Engle and
Chapter V: Discussions
102
Gustin, 2002) and substrates associated with hydrothermal systems (Engle
et al., 2006).
Calculated fluxes of Hg release from sediments result to be one order of
magnitude higher than reported total sea-air mercury evasion flux,
suggesting that the a good part of Hg released from sediments keep on
recycle in the biogeochemical cycle involving sediment, seawater and biotic
compartment. The total amount of each compartment remain unclear.
5.4 Mass-balance
Recently Sprovieri et al. (2011) and Fantozzi et al. (2013) have argued on a
potential Hg export from Augusta Bay to the Eastern Mediterranean
seawater, as a result of the measured 3D circulation system. Rajar et al.,
2007 reported a total input of Hg from anthropogenic sources to the
Mediterranean Sea of ~12.5 kmol y-1
excluding the Augusta Bay
contribution.
Here, we extend the potential role of Augusta Bay as an Hg point source for
the open ocean also considering the significant transfer of pollutants by
pelagic fishes moving between the inner and external part of the bay. This
implies potential effects on the food web of the surrounding area as already
reported by several authors studying marine systems (Riisgard and Hansen,
1990; Futter, 1994; Jarman et al., 1996; Atwell et al., 1998). This hypothesis
is supported by the high mean HgT concentrations measured in pelagic
species caught outside the bay. Indeed, pelagic species prefer to inhabit
warmer coastal seawaters during their first life stages (Basilone et al., 2011),
but they generally move in deeper waters during the older stages
(Wirszubski, 1953; Schneider, 1990; Whitehead, 1990), thus, representing a
significant and potential vehicle of contaminants to the deep marine food
web.
This evidence corroborates our hypothesis of a potential Hg export through
the food web, from Augusta Bay to the surrounding area. Moreover the high
Hg content measured in seawater collected outside the bay, confirmed the
hypothesis of transport of Hg from Augusta harbor to the open sea
representing a vehicle of contamination for the entire Mediterranean basin
through the complex circulating currents affecting the western Ionian.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
103
The above evidences definitively called for a more detailed information on
the role played by the Augusta Bay on a large scale in order to assess the net
outflow of Hg through the offshore waters.
A box model that calculates the budget of HgT in seawater and sediments of
the Augusta area is here proposed. Sprovieri et al., 2011 considered the
Augusta Bay as an ‘environmental compartment with well-defined borders
and constrained hydrodynamics” and reported an average output of water
mass of ~2.34 x 1013
kg y-1
. The mass balance of HgT in the basin is
calculated at the steady-state by the following equation:
I + A + AD + R = O + D + V
Where:
I represents the total Hg influx from the surface Mediterranean seawater;
A represents the inputs of dissolved Hg from anthropogenic activities
AD corresponds to the atmospheric Hg deposition
R is the Hg re-suspension/release from sediments
O is the total Hg outflow from the basin
D corresponds to sedimentary deposition and burial
V is a term accounting for evasion of HgT to the atmosphere (Fig. 35).
All the budgets and parameters described below are summarised in Table
16. Values of I and A are calculated using the HgT concentrations reported
by Kotnik et al., 2007 (1.46 ±0.62 and 1.21 ±0.32 pM in summer and spring,
respectively), corresponding to an average input of Hg in the Augusta basin
(I) of about 8.48 x10-7
kmol y-1
from the Ionian Sea (Fig. 38).
Values of anthropogenic HgT inputs to the basin (A) (0.018 t in water for
the year 2005), are reported in the European pollutant emission register.
Values of AD and V are those reported by Bagnato et al, 2013 performing a
short-term (1 yr) monitoring study on Hg distribution and evasional flux in
the atmospheric compartment of the Augusta Bay.
The calculated V (1.99 x10-2
kmol y-1
) result to be one order of magnitude
higher respect of data used by Sprovieri et al., 2011, while the AD value
(4.19x 10-3
kmol y-1
) results double. The applied benthic chamber
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
104
experiment offered the unique opportunity to measure the authentic release
of Hg from Augusta Bay sediment to seawater over time (R).
The R parameter was calculated by extending the in situ benthic flux to the
whole Augusta area by the application of the Voronoi polygons methods,
dividing space into a number of regions, represented by the investigate site.
The method define individual areas of influence around each of a set of
points. Polygons are mathematically defined by the perpendicular bisectors
of the lines between all points (Fig. 38).
a)
b)
Figure 38: Partition of the entire Augusta basin in Voronoi polygons,, each relative to one
station of measurement for September 2011 (a) and June 2012(b)benthic chamber
experiments.
Chapter V: Discussions
105
By extending the punctual fluxes to the entire Augusta area, a total amount
of 0.38 Kmol y-1
has been estimated (Fig. 39). As Sprovieri et al., 2011, on
consider null D parameter.
The calculated output of HgT from the Augusta basin to Ionian surface
waters (O) corresponds to an average of 1.29 kmol y-1
(Fig. 39), more than
one order of magnitude respect of those measure by indirect calculation
reported by Sprovieri et al., 2011.
On the basis of the available hydrodynamic data, Sprovieri et al., 2011
calculate a residence time for the bottom water of the Augusta of ~10
months and therefore it is assumable that the HgT released from sediments
to the bottom seawater of the basin is flushed on a yearly basis and outflows
to the eastern coast of Sicily (Sprovieri et al., 2011).
Table 16: Parameters (sources and sinks) of HgT used to calculate the mass-balance of
HgT in the Augusta basin.
Unit Ia A
b ADc V
c Rd D O
Kmol y-1 1.70 x 10-7 1.80 x 10-2 4.19x 10-3 1.99 x10-2 0.38 - 1.29 a Kotnik et al. (2007)
b From the EPER database
c Bagnato et al., 2013
d this work
Figure 39: Conceptual scheme for mass-balance calculation of HgT in the Augusta basin.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
106
The most critical point related to the role played by the Augusta Bay is the
potential large scale effects associated to the outflow of waters intercepting
surface meso-scale ocean circulation (Sprovieri et al., 2011). Indeed, water
mass outflow from the Augusta Bay is immediately intercepted by the
Atlantic Ionian Stream (AIS) flowing over the continental shelf off the
southern Sicilian coast, due to the general thermohaline circulation in the
upper, intermediate and deep layers, creating an anticyclone circulation.
Furthermore the narrow continental margin off the Augusta coast,
associated to steep slope and several gullies (Sprovieri et al., 2011), creates
preferential transfer routes for polluted sediments to the Mediterranean
basin.
The recalculated input provided by the Augusta Bay (1.29 Kmol y-1
) have to
be added as point sources to the previous mass balance estimate. By adding
the estimated Hg input given by the Augusta basin to the total mercury input
reported by Rajar et al., 2007 for the Mediterranean sea, the total output
rises to 319.11 kmol y-1
(Fig. 40).
Figure 40: HgT flows calculated for the Mediterranean Sea with the Augusta source point
(Rajar et al., 2007 and Sprovieri et al., 2011 modified).
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
107
5.5 REEs distribution in seawater
The use of REEs distribution as pollution tracer in environmental medium is
due to the following evidences:
Due to their unique chemical features, REE represents, in all
likelihood, the best geochemical instrument for investigate processes
realizing at the interface among different means, both of organic or
inorganic origins (Bau & Dulski, 1996a; Bau, 1999; Takahashi et al.,
2005; 2007; 2010).
The activities related to oil refining industry produces a
characteristic geochemical signature that REE distributions are able
to record as positive anomalies of light REE (LREE), both in
dissolved and in particulate phases suspended in atmosphere and
hydrosphere (Olmez et al., 1991; Kulkarni et al., 2007; Moreno et
al., 2008; Censi et al., 2011).
The geochemical behavior of REE and their speciation in the
dissolved phase are strongly influenced by the organic matter
content. This evidence has also been recently demonstrated in
coastal marine environment and constitutes a tracer of anthropic
pressure in coastal water (Censi et al., 2010).
Various work has addressed the use of REEs in investigating the
environmental impact of human activity and demonstrated that the REE
natural distribution in water, soil, and sediment from densely industrialized
and populated regions can be altered by anthropogenic influences (i.e. Bau
and Dulski, 1996). This is a very important topic in the study of coastal
areas, where strong anthropogenic pressure occurs such in as the
Mediterranean Sea (Elbaz-Poulichet et al., 2002; Di Leonardo 2009; Censi
et al., 2010). Particularly in the Augusta Bay the combination of industrial,
agricultural and urban effluents, as well as dry and wet deposition, play a
determining role on the evolutionary process of chemical characteristics in
the Augusta marine system. Features of shale-normalised REE pattern in
seawater are here used as proxy to recognize sources of these elements and
processes involving them along the water column.
The main geochemical process highlighted by the REEs distribution in the
examined samples, is the failure scavenging of Ce from the dissolved phase
to the suspended particulate matter. Cerium anomaly in natural waters is
related to the mechanism of oxidative scavenging of tetravalent cerium by
Chapter V: Discussions
108
iron and/or manganese oxides. Ce abatement, or its complete disappearance
has been recovered in continental water of varies origins and results
attributable to organic matter content (Dia et al., 2000). Pourret et al., 2008
showed that above pH 8.2, 8.6 or 8.7, Ce(III) is readily oxidized into
Ce(IV), which is then preferentially adsorbed onto humic acids, resulting in
the development of a negative Ce anomaly in the “truly” dissolved part of
the solution, and a complementary positive anomaly occurs in the organic
colloidal fraction.
In the reported dataset, the absence of Ce/Ce*<1 in dissolved phases is
probably due to high concentration of dissolved organic. Ultra-filtration
allows to physically separate inorganic to organic phases, permitting to
discern the negative Ce anomaly (Pourret et al., 2008). Although this
behavior has been detected only in laboratory studies (Davranche et al.,
2004, Pourret et al., 2008) and confirmed for marine environment in a
limited number of cases (Censi et al., 2010), data collected in Augusta Bay
suggest this result could be due to the above described mechanism. The
known anthropogenic pressure in the study area and the presence of
petrochemical plant, the Gd anomaly can be ascribed to zeolites doped with
REEs, which have been used as fluid-cracking catalysts in petroleum
industry since the 1960s (Swift et al., 1979; Kulkarni et al., 2006).
The contemporary and correlated presence of positive Gd and Ce anomalies
in seawater, let suppose a common origin. Data confirmed that Ce anomaly
is sensible to dissolved organic matter, more than redox condition, in sample
water submitted to simple filtration. The REE distribution along the water
column suggests that the high dissolved organic matter increase the REE
solubility from sediments in which they are concentrated because of
industrial activities. Waste discharge of industrial wastewater, together with
human non industrial wastewater, created ideal condition for an increasing
of REE in dissolved phases.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
109
5.6 Application of Stable Mercury Isotopes to
biogeochemistry
The application of mercury (Hg) stable isotopes to problems in
environmental chemistry is rapidly expanding field of investigation. The
success of such approach strongly depends on two factors: first, different Hg
sources must have analytically discernible Hg isotope signatures; second,
some Hg geochemistry transformation have a known isotopic signature. The
combination of mass-dependent (MDF) and mass-independent (MIF)
fractionation provides a new isotopic fingerprint of mercury sources and
processes into the environment. Here the Hg isotopic investigation on
fishes, sediments and hairs permitted to define sources of Hg and the
biogeochemical processes realized in Augusta Bay. A direct positive
relationship (R2= 0.92) has be found by plotting Δ
201Hg vs. Δ
199 values of
the entire dataset, included both fish, box-corer sediments and human hair of
local population (Fig. 41). This clear reflect the crucial link between the
main Hg source (sediment) and the last receptor, the man, assuming Hg
primary by mean of fish consumption.
Δ199Hg = 0.0095+1.0808*x
Fish
Sediment
Hair
-0.6 -0.4 -0.2 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6
Δ201Hg
-0.2
0.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
1.6
1.8
Δ1
99H
g
Δ201Hg:Δ199Hg:
r2 = 0.9222;
r = 0.9603;
p = 00.0000;
y = 0.0095 + 1.0808*x
Fig. 41: Plot of data for Δ
201Hg vs. Δ
199Hg of the Augusta Bay coastal sediments, fish
muscle tissues and human hair of local population (r2=0.92; slope: 1.08±0.009).
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
110
5.6.1 Hg isotopes fractioning in fishes
Variations in Δ199
Hg and Δ201
Hg in fish tissues resulted from photochemical
degradation of MeHg in the water column prior to its entry into aquatic food
chain and the Δ199
Hg /Δ201
Hg ratio might be diagnostic of reduction of Hg
(II) vs. MeHg. (Blum and Bergquist, 2007, Sherman and Blum, 2013).
Indeed empirical observations of Δ199
Hg/Δ201
Hg slopes in natural materials
show values of ~1.0 for processes that involve Hg(II) photoreduction
(Biswas et al., 2008; Ghosh et al., 2008; Carignan et al., 2009; Sherman et
al., 2010) whereas values in aquatic foodwebs, which uniquely involve
MeHg, display slopes of ~1.2–1.3 (Bergquist and Blum 2007; Gantner et al.,
2009; Laffont et al., 2010; Senn et al., 2010). Here the positive MIF trend,
and the Δ199
vs Δ201
correlation (slope = 1 ± 0.03) clear testify Hg
fractioning appended during photochemical reaction of Hg prior of the
intake in the marine food web. MeHg display more MIF compared to Hg
(II) and the ratio Δ201
Hg vs δ202
Hg is predictive of photos or demethylation
reduction. The calculated Δ201
Hg / δ202
Hg for fishes results to be 1.7,
confirming the most important reaction causes fractioning is the Hg(II)
reduction prior of the intake in the aquatic food web.
Moreover MIF fractionate otherwise according to fish ecology. Senn et al.
(2010) reported for coastal species in the Gulf of Mexico distinct and non-
overlapping values of δ202
Hg and Δ201
Hg compared to oceanic specie.
Difference in Δ201
Hg between sediments and fish results in enrichment of
0.7‰ in fishes, while for Δ199
Hg the difference result the 0.67 ‰.
The clear increase in MIF fractionation (both Δ199
Hg and Δ201
Hg) in pelagic
fishes could be explained by their living habitat, more surface respect of
demersal and benthic species, where penetrating sunlight promotes the
processes of Hg photoreduction in the water column. On the other hand the
higher positive δ202
Hg values measured in the benthic and demersal
specimens could be due to microbial degradation of MeHg before entry into
the food web.
The main pathway drive this fractioning can be summarizing as:
Hg (II) Hg(0)
Chapter V: Discussions
111
5.6.2 Hg isotopes fractioning in sediments
MDF fractionation occurs during biotical methylation processes in
sediments. Studying fermentative methylation by sulphate reducing
bacteria, Rodriguez-Gonzalez et al., 2009, demonstrate that the MeHg
product had lower δ202
Hg than the Hg(II) starting material (-MDF) and also
found no evidence for biotic MIF during methylation, contrary to Jackson et
al. (2008) that argued that MIF of Hg occurred during microbial Hg
methylation.
By plotting δ202
Hg vs Δ201
Hg for the entire sediment core dataset, on obtain
a peculiar trend (Fig. 35), ascribable, according to the scheme proposed by
Blum (2011) and Rodriguez-Gonzales et al., (2009), to biological mediated
MeHg production. Moreover processes of Hg methylation in sediments
could be drive the shift in δ202
Hg measured in the south and most
contaminated Augusta Bay.
The highest Δ199
Hg and lower δ202
Hg recorded in the south area let
hypothesize the higher microbial activity in the most contaminated area.
This geographic pattern in δ202
Hg and Δ199
Hg values suggests that the
sources of Hg to the sediment are locally controlled.
The gradually changing spatial pattern in sediment Hg isotopes is consistent
with sediment Hg contamination derived from chlor-alkali plant, probably
due to a stronger microbial activity in the southern Bay. The main pathway
drive this fractioning can be summarizing as:
Hg(II) MeHg = microorganism
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
112
5.6.3 Hair Hg isotopic composition
The importance of the Hg/DOC ratio and the effect of ligands on Hg isotope
fractionation, has been tested by the exploration of photoreduction of Hg(II)
in a series of laboratory experiments (Zheng and Hintelmann 2009, 2010b).
Laboratory experiment conducted by Bergquist and Blum, 2007, used
natural sunlight to reduce Hg(II) and MeHg from aqueous solutions in the
presence of DOC, demonstrated Δ199
Hg /Δ201
Hg ratio was 1.00 for Hg(II)
reduction and 1.31 for MeHg reduction.
The authors suggested that this range represented differing behavior of Hg
bonded predominantly to reduced sulfur groups at low Hg/DOC and to
oxygen groups at higher Hg/DOC. The analysed hair samples display
positive MIF with a slope Δ201
Hg vs. Δ199
Hg of 1.2 clearly indicated
processes of MeHg-DOC reduction due to sunlight.
The main pathway drive this fractioning can be summarizing as:
MeHg-DOC Hg(0)
5.6.4 Sources of Hg for fishes
Several studies have explored the use of Hg isotopes to “fingerprint” point
sources of Hg from historic Hg mines and trace them into the surrounding
environment (Gehrke et al., 2011c) .
If Hg transformations do shift original source isotope signatures, then under
certain conditions the shift may prove to be constant. Examples are the
observed shifts in δ202
Hg by 0.6‰ between sediment Hg and fish Hg in San
Francisco Bay reported by Gehrke et al., 2011c.
Here on report δ202
Hg values analogous for both fishes and sediment (mean
δ202
Hg=-0.39 ‰), doubtless confirming the role of sediments as source of
Hg for the marine biotic compartment.
The support at the hypothesis of sediments as sources of Hg for fishes, is
also testimony by the high-quality relation (r2=0.90) obtained by plotting
δ202
Hg vs. Δ201
Hg of both fishes and sediments (Fig. 42).
The higher MIF displayed by fishes had been discussed above. MeHg
photochemically demethylation in the coastal ecosystem, prior to being
incorporated into the oceanic food web has been supported by experimental
results reported Bergquist and Blum (2007), observing MIF in oceanic fish.
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
113
Fish
Sediment -0.6 -0.4 -0.2 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6
Δ201Hg
-0.2
0.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
1.6
1.8
Δ199
Hg
Δ201Hg :Δ199Hg
r2 = 0.9013;
r = 0.9493;
p = 00.0000;
y = -0.0003 + 0.9723*x
Fig. 42: Plot of data for Δ201
Hg vs. Δ199
Hg for the Augusta Bay coastal sediments and fish
muscle tissues.
5.6.5 Sources of Hg for human
The most important pathway for human exposure to Hg is via the
consumption of marine food (Sunderland 2007). Compared to blood and
urine, hair, because of the stable Hg forms, is most suitable for monitoring
study in polluted site (Gellein et al., 2008; Dongarrà et al., 2010). Here on
investigate the Hg isotopic variations between fish tissue and human hair as
a first step in evaluating Hg isotopes as a metabolic process and/or Hg
exposure.
Here the meaningful relationship found by plotting Δ201
Hg vs. Δ199
Hg (R2=
0.84; slope: 1.09, Fig. 43), demonstrate the familiarity between the Hg
content in fishes and hairs.
Examples are the observed shift of 2‰ in δ202
Hg between fish and the scalp
hair of human consumers of fish has been reported in San Francisco Bay
(Laffont et al., 2010). For Augusta Bay compared to fish (mean δ202
Hg= -
0.39), human hair (mean δ202
Hg= 1.71) shift for 2.1‰, while mean MIF
difference are negligible (0.3‰ for Δ201
Hg e 0.5 ‰ for Δ199
Hg), suggesting
Chapter V: Discussions
114
that the anomalies act as conservative source tracers between upper trophic
levels (men) of the tropical food chain (fishes), in good agreement with
results reported by Laffont et al., (2010).
Δ199
Hg = -0.0016+1.0943*x
Fish
Hair -0.2 0.0 0.2 0.4 0.6 0.8 1.0 1.2 1.4 1.6
Δ201
Hg
-0.2
0.0
0.2
0.4
0.6
0.8
1.0
1.2
1.4
1.6
1.8
Δ199H
g
Δ201
Hg:Δ199
Hg:
r2 = 0.8433;
r = 0.9183;
p = 00.0000;
y = -0.0016 + 1.0943*x
Fig. 43: Plot of data for Δ
201Hg vs. Δ
199Hg for fishes and human hair samples.
Although some MDF probably also occurs during trophic transfer of MeHg
through the food web (Laffont et al., 2010, Das et al., 2009), trophic transfer
could also push δ202
Hg values to higher values. Laffont et al., 2010 justified
variation of ~ +2 ‰ to MMHg human metabolism, i.e. excretion in faeces
of light isotopes of Hg or blood-hair transfer, that determine enrichment of
heavier Hg isotopes (δ202
Hg).
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
115
Chapter VI: Conclusions
Chapter VI:
Conclusions
116
6. CONCLUSIONS
Mercury has an extremely complex cycle in the Earth’s ecosystems and the
environmental bodies can be both active sink and source of Hg. The here
reported multidisciplinary work permits to put a complete view on the
biogeochemical behavior of Hg in a coastal marine environment. In this
contest the Augusta Bay represent an ideal natural lab for study on the
biogeochemical cycle of Hg in complex systems, because of relevant and
uncontrolled discharges of Hg from the chlor-alkali plant. Analyses of Hg
content in fishes, sediments, seawaters and atmosphere of the Augusta Bay
allowed to investigate to the major actors of the cycle. Moreover fluxes
studies at the interfaces sediment/seawater/air, performed by means of an in
situ benthic chamber and of a dynamic accumulation chamber, allowed to
accurately assess the fraction of Hg released from sediments and that
escaped in atmosphere. Finally the creation of an integrated model on the
biogeochemical cycle of mercury permitted to estimate the Hg contribution
made by the Augusta Bay as point source for the Mediterranean basin, by
means of the complex of water circulation involving the eastern Sicilian
coast.
The main conclusion can be sensitized in the following sentences:
Sediments of Augusta Bay played a key role as source of Hg for the
surrounding environment. A still active Hg release mechanism from the
polluted sediments to the marine environment is testified by several
evidences:
The highest Hg content, in both fishes, sediments and seawater samples, had been
recorded in the southern Augusta Bay, where waste spillage from chlor-alkali plant
occurred.
The HgT content measured in fish tissues show an increasing trend along the
habitat depth, with highest values measured in benthic species and the lowest in
pelagic ones.
THg and DHg content in seawater shows a clear trend on the vertical, with
concentration increasing near the bottom and reducing in surface water.
The total DHg content at the sediment/seawater interface increases with time.
Fishes and sediments had the same isotopic signature, in terms of mass dependent
fractionation (δ202
Hg).
M. Bonsignore: “The biogeochemical cycle of Hg in the Augusta Bay
117
Active processes of Hg microbial methylation in sediments are testified by Δ201
Hg/
δ202
Hg ratio in sediments.
The Augusta Bay represents a point sources for Hg to Mediterranean
basin. The vehicle of contamination is favorite by the complex circulating
currents affecting the western Ionian, by the closeness with the steep
continental slope (part of the Malta escarpment) and by the presence of
several gullies, dropping to the deep end of the Ionian basin. This
conclusion is supported thought the following evidences:
High contamination of pelagic species measured in the external zone of the bay
confirmed the role of the Augusta marine environment as a potential Hg source for
the surrounding area and underscores the crucial risk associated with contaminant
transfer from the semi-enclosed basin to the open sea by means of biotic transfer.
Unexpected THg concentrations have been measured in seawater outside the bay;
The mass balance calculation estimated a HgT output from the Augusta basin to
Ionian surface waters (O) corresponding to an average of 1.29 kmol y
-.
The REE distribution along the water column suggests that the high
dissolved organic matter created ideal condition for an increasing of REE in
dissolved phases, much to hide the negative Ce anomaly usually recorded in
the oligothropic water. Gd anomaly, expressed as Gd/Gd*>1, let’s suppose
significant contributions arising from the petrochemical plant.
Chapter VI: Conclusions
118
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Acknowledgements
One of the joys of completion is to look over the journey past and remember
all the friends and family who have helped and supported me along this
challenging and rewarding road.
I would not have contemplated this road if not for my family, gave me trust
and confidence.
I would like to express my heartfelt gratitude to doctor Mario Sprovieri,
who was for me not only mentor but also ethics guide. I could not have
asked for better model of inspiration, tenacity and love to research.
A dear thanks goes to Salvatore Mazzola, for giving me his warmly
welcomed at the IAMC-CNR Institute.
My gratitude goes to the professor Paolo Censi, for his clef contribution in
REE study.
As member of the biogeochemical team of the IAMC-CNR Institute, I have
been surrounded by wonderful colleagues and someone friends. Elvira
Oliveri and Daniela Salvagio Manta, loving guides and fellow of a
unforgettable rout. Thanks for being such more than work colleagues..this
achievement is also yours!!!
Special thanks to Emanuela Bagnato which with her manuscript has made a
key contribution for the whole work.
Many other people helped me in some way, supporting me on work and in
life: Gualtiero Basilone, Simona Genovese, Marco Barra, Francesca
Bulfamante, Angelo Bonanno, Giorgio Tranchida, Vincenzo Di Stefano,
Carlo Patti, Francesco Feliciotto, Enza Quinci, Anna Traina, Rosalia
Ferreri, Marianna Del Core.
I extend warm greetings to Professor Andrea Marchetti and to Caterina
Durante (Unimore University, Modena) that permitted me to explore the Hg
isotopes world and provided a rich and fertile environment to study and
probe new ideas. Thank you to Stella Tamburrino, my mate in this nice
experience.
Acknowledgments
133
I would also like to thank my reviews, doctor Stefano Covelli (University of
Trieste, Italy) and professor Pierpaolo Zuddas (Paris-Sorbonne Institut),
and my examiners, who provided encouraging and constructive feedback. I
am grateful for their thoughtful and detailed comments. Thank you for
helping to shape and guide the direction of the work with your careful and
instructive comments.
Thank you also to my dear friends, Claudia and Rosario who always hold
up me with their undying affection.
And last, but not least, to Filippo, who shares my passions, thank you for
keep alive my dreams.
All of you have been like surrogate families, bearing the brunt of the
frustrations, and sharing in the joy of the successes.
Acknowledgments
134
Attachment I
135
Ma
MAa
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heroinm2draow
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0h
ercury in fishes from Augusta Bay (southern Italy): Risk assessment
Food and Chemical Toxicology 56 (2013) 184–194
Contents lists available at SciVerse ScienceDirect
Food and Chemical Toxicology
journal homepage: www.elsevier .com/locate/ foodchemtox
b, E. O
NR, Via deNR, Calata
b s t r
r studyal benthisemi-enc
nd health implication
. Bonsignore a, D. Salvagio Manta. Traina a, S. Mazzola a
Institute for Coastal and Marine Environment (IAMC) – CInstitute for Coastal and Marine Environment (IAMC) – C
r t i c l e i n f o
rticle history:eceived 15 December 2012ccepted 14 February 2013
a
Ouera
vailable online 24 February 2013
eywords:ercury
charge of tregard to sResults sugment. Also
011) reported high-resolution maps of HgT distruperficial sediments collected in 2005, highlightin
278-6915/$ - see front matter � 2013 Elsevier Ltd. All rights resettp://dx.doi.org/10.1016/j.fct.2013.02.025
⇑ Corresponding author. Tel.: +39 092440670; fax: +39 0924404E-mail address: [email protected] (M. Sprovieri).
liveri a, M. Sprovieri a,⇑, G. Basilone a, A. Bonanno a, F. Falco a,
l Mare, 3, 91021 Torretta Granitola, Campobello di Mazara (TP), ItalyPorta di Massa, 80100 Naples, Italy
a c t
reports on the total mercury (HgT) concentrations measured in the muscles and livers of sev-c, demersal and pelagic fish species caught inside and outside of Augusta Bay (southern Italy),losed marine area, highly contaminated by the uncontrolled (since the 1950s to 1978s) dis-he largest European petrochemical plant. Mercury levels in fish tissues are discussed with
pecific habitat, size and/or age of the specimens and HgT distribution in the bottom sediments.gest a still active Hg release mechanism from the polluted sediments to the marine environ-
, the high HgT concentrations measured in fishes caught in the external area of the bay imply a
ishesollution effectioaccumulationpotential role of Augusta Bay as a pollutant source for the Mediterranean ecosystem. Finally, values ofhazard target quotient (THQ) and estimated weekly intake (EWI) demonstrate that consumption of fishes
isk fpecik for
ntratingg toby
ly c) aniotindatedl Hicrot al
rt os frse
lsbet ah wspecurre
oxicity caught inside the bay represents a serious rcaught from the external area of the bay, esbe represent a significant component of ris
. Introduction
The city of Augusta, located in the SE of Sicily (southern Italy),as experienced an important industrialisation phase since thearly 60s. This has led to the creation of several chemical and pet-ochemical plants and oil refineries resulting in a severe pollutionf the surrounding environment. In particular, the petrochemicaldustry in Augusta Bay is one of the largest in Europe with theost important chlor-alkali plant in Italy (Le Donne and Ciafani,
008). Its activity started in 1958 and stopped in 2005, with pro-uction of chlorine and caustic soda by electrolysis of sodium chlo-ide aqueous solution in electrolytic cells with a graphite anodend metallic mercury cathode. Uncontrolled chemical dischargef Hg occurred in the Augusta Bay until 1978, when restrictionsere imposed by the Italian legislation.
In the last decade, several studies have provided detailed infor-ation on the pollution levels and risks for human health of resi-
ent populations of Augusta Bay (ICRAM, 2005; Ausili et al.,008; Di Leonardo et al., 2007, 2008; ENVIRON International Team,008; Ficco et al., 2009; Sprovieri et al., 2011). Sprovieri et al.
high conceand speculaexporting Hintercepteddata recentTeam (2008from the abpredators arisks associToxicologicamullet by mage (Ausili e(2012) repolis specimen
Fish foodhumans (Ho
Renzoniintake of fistoxic risk, eepisode occ
ibution fromg extremely
mata Bay (Japaand teratogenicfish (De Flora etform, able to ininactivation of
rved.
45.
or human health. Also, data indicate that intake of fishesally for that concern demersal and benthic species, could
the local population.� 2013 Elsevier Ltd. All rights reserved.
tions (ranging between 0.1 and 527.3 mg kg�1)on the key role that Augusta Bay could play in
the Mediterranean Sea, as an effect of the outflowthe Levantine Intermediate Waters (LIWs). Also,
ollected by ICRAM (2008), ENVIRON Internationald Ausili et al. (2008), demonstrated HgT transfer
c system (sediments and seawater) to fishes (topfilter-feeders) and documented significant health
with the consumption of fish caught in the area.g effects were also evaluated on mussels and rednuclei (MN) studies, which documented DNA dam-., 2008 and ICRAM, 2008). Finally, Tomasello et al.
n DNA genotoxic and oxidative damages in Coris ju-om Augusta Bay.ems to constitute the main route of Hg uptake foreek et al., 1996; Nakagawa et al., 1997).l. (1998) demonstrated that long-term and frequentith high Hg levels is statistically associated with aially in pregnant women. A sad, famous poisoningd in the 1950s among people living around Mina-n), showing the irreversible neurological damageeffects due to consummation of Hg-contaminatedal., 1994). Methylmercury (MeHg) is the most toxic
terfere with thiol metabolism, causing inhibition or
proteins containing thiol ligands and ultimatelyleNifas19et
foul20tiofhe
Auseat
2.
2.1
gubodedepefopomiti(ospinbounanCa
2.2
w(1atmgrmcean
2.3
mac0.1tew(Tcan =m
2.4
lat
m
TH
xposvalen2005ce dotermrage(365
as cance dn parformenry,ted wtimeseigh
an THng th. 73/ratel
l fea
caug2 penspel
spectotaies,, Mu
(201
ading to mitotic disturbances (Das et al., 1982; Elhassani, 1983).umerous recent studies indeed have concluded that the majority,not all, of the Hg that is bioaccumulated through the food chain is
MeHg (Winfrey and Rudd, 1990; Mason and Fitzgerald, 1990,91; Gilmour and Henry, 1991; Horvat et al., 1999; Carbonellal., 2009).High mortality rates, statistical high frequency of neonatal mal-
rmations and cancerous diseases reported for resident pop-ations around Augusta Bay (Martuzzi et al., 2006; Bianchi et al.,04, 2006; Fano et al., 2005, 2006; Madeddu et al., 2006) defini-
vely calls for more detailed exploration and definitive assessmentthe role played by the intake of Hg-contaminated fish on thealth of the consumers.In this work, we aim to explore the effects of HgT pollution ingusta Bay on the fish compartment, inside and outside the
mi-enclosed area, and to assess the potential health risks associ-ed with the consumption of contaminated fish.
Materials and methods
. Sampling
Four different sampling sites were selected: two inside, and two outside of Au-sta Bay (Fig. 1). Sampling outside the bay was performed during May 2001, onard of the N/O ‘‘Dallaporta’’, by means of a mid-water trawl-net at 50–100 m ofpth in two sampling areas, in front of the Scirocco inlet (300-m wide and 13-mep), and the Levante inlet (400-m wide and 40-m deep) (Fig. 1: C1, C2). Mainlylagic fish specimens were caught (Table 1). Sampling inside the bay was per-rmed during May 2012 by means of a fishing boat equipped with a gillnet wall,sitioned at the bottom (mean depth = 20–25 m) (Fig. 1: C3, C4). Several speci-ens of benthic and demersal fishes were collected. From the two sampling activ-
where EF is e(70 years), equison/day) (FAO,USEPA’s referendaily intake deWAB is the aveno carcinogens
The THQ wUS-EPA’s refereWHO (THQb). Iits methylatedGilmour and H
The estimacentration (C)average body w
Finally, meAlso, consideri2007 (Order Noprocessed sepa
3. Results
3.1. Biologica
The fishconsisted ofwhile specimsisted of 10321 differentspecies andcaught specBoops boops
M. Bonsignore et al. / Food and Chemical Toxicology 56
es, a total of 227 fish specimens were collected: 107 from mid-water sampling Mediterranean S). Ospecof th(Stre
rcur
rcurd bephiclate
utside the bay) and 120 from bottom-water sampling (inside the bay). Moreover,ecimens of Engraulis encrasicolus (n = 38) were caught from the unpolluted mar-e area of Marsala (western Sicily) (Fig. 1), during July 2001, on board of a fishingat equipped with a purse seine net. After collection, fishes were stored at �20 �Ctil biological and chemical analyses were performed in the laboratories of biologyd biogeochemistry at the Institute for Coastal and Marine Environment (CNR) ofpo Granitola.
. Biological data and tissue collection
The total length (TL) of each specimen was measured. Muscle and liver tissuesere collected from each organism, using plastic materials cleaned with HNO3
0%) and MilliQ water, in order to avoid Hg contamination. Tissues were stored
(Irepa, 2010called alienical speciesranean Sea
3.2. Total me
Total medemersal anBay, are gravalues calcu
�20 �C until analysis. Otoliths were extracted from anchovy and sardine speci-ens for age determination. Readings and interpretation of otolith increment
ative data fromarsa
coin
ins hi
Table th(Hgmenint i
ananlg
sent-out
andcorponc) anhe laugh.570ed i
owths were carried out by transmitted visible lights based on higher-resolutionicroscopy (20–25�magnification) (Campana et al., 1987; Nielsen, 1992). The pro-dure adopted for European anchovy age determination follow Uriarte et al. (2007)d La Mesa et al. (2009).
. Chemical analyses
Total mercury concentrations (HgT) in tissues were measured using a directercury analyser (Milestone_DMA-80), atomic absorption spectrophotometer,cording to analytical procedures reported in EPA 7473. Briefly, approximately
g of fresh tissue was loaded in nickel boats and transferred to the DMA-80 sys-m. In order to minimise contamination risks, acid-cleaned laboratory materialsere used during sample preparation and analyses. A Reference Standard MaterialORT-2; HgT certificate value = 0.27 ± 0.06 lg g�1) was analysed to assess analyti-l accuracy (estimated to be �3%) and precision (routinely better than 4%; RSD%,
3). Finally, duplicated samples (about 20% of the total number of samples) wereeasured to estimate reproducibility, which resulted in better than 7%.
. THQ and EWI calculation
Target hazard quotient (THQ) and estimated weekly intake (EWI) were calcu-ed for muscles of fishes caught inside and outside the bay.
The target hazard quotient was calculated according to the US EPA (1989)ethod and it is described by the following equation:
Q ¼ EF� ED � FIR � C� �
� 10�3
vies from MMercury
2.709 lg g�1
9.720 lg g�1
1.5 to 6 timespecimens (caught insiddus vulgarisand a speci(extreme po(HgT = 9.720(Table 2)(HgT = 2.638levels reprehighest nonside the bayparticular, SHgT mean crespectively(Table 2). Tspecimens cand 0.029–0ues measur
RFD�WAB� TA comparable (Tab
ure frequency (365 days/year); ED is the exposure durationt to the average lifetime; FIR is the food ingestion rate (36 g/per-); C is the metal concentration in seafood (lg g�1); RFD is these (0.1 lg Hg kg bw�1 d�1) (http://cfpub.epa.gov) or acceptableined by WHO (0.23 lg Hg kg bw�1 d�1) (http://apps.who.int);body weight (60 kg), and TA is the average exposure time fordays/year � ED).
lculated for all the studied species in the Augusta Bay using theose (THQa) and the acceptable daily intake determined by the
ticular, we assumed that the measured mercury is integrally in(Winfrey and Rudd, 1990; Mason and Fitzgerald, 1990, 1991;1991; Horvat et al., 1999; Carbonell et al., 2009).eekly intake (EWI) was calculated by multiplying the HgT con-by the weekly dietary intakes (FIR � 7) and reporting to the
t (WAB).Q and EWI values were calculated for each studied species.
at fishing activity within the bay has been interdicted since07), data relative to fishes from inside and outside the bay werey.
tures
ht from bottom-water sampling (inside the bay)elagic, 106 demersal and 16 benthic specimens,from mid-water sampling (outside the bay), con-
agic, 3 demersal and 1 benthic (Table 1). A total ofies were recognised. The number of specimens per
l length ranges are shown in Table 2. Almost all thein particular, E. encrasicolus, Sardina pilchardus,llus barbatus and Illex coindetii, are typical of theea and are commercially relevant to Italian fishingnly one specimen was found to belong to a so-ies, specifically Sphyraena sphyraena. This is a typ-e tropical seas, today present also in the Mediter-ftaris and Zenetos, 2006).
y concentrations (HgT)
y concentrations measured in tissues from pelagic,nthic fishes, caught inside and outside of Augustaally summarised in Fig. 2a and b. Mercury mean
d for each species, together with available compar-the literature and HgT content measured in ancho-la, are presented in Table 2.ncentrations ranged between 0.021 and
muscles (Fig. 2a) and between 0.029 andlivers (Fig. 2b). The HgT content in liver is fromgher than that measured in muscles from the samee 2). The highest HgT values were found in speciese bay: 2 demersal specimens, a specimen of Diplo-T in liver = 4.979 lg g�1) (extreme point in Fig. 2b)
of Serranus scriba (HgT in muscle = 2.709 lg g�1)n Fig. 2a), a large pelagic specimen of S. sphyraenad 2.269 lg g�1 in liver and muscle, respectively)d a benthic specimen of Murena helenag�1 in muscle) (Table 2). However, these very highoutliers of the whole dataset (Fig. 2a and b). The
lier values refer once again to specimens caught in-specifically to benthic species (Fig. 2a and b). In
anea scrofa and Scorpanea notata show the highestentrations for both liver (1.638 and 2.339 lg g�1,d muscle (1.082 and 1.341 lg g�1, respectively)owest non-outlier ranges were found in pelagict outside the bay (0.021–0.167 lg g�1 for muscles8 for livers) (Fig. 2a and b), and the HgT mean val-n the different studied species are substantially
3) 184–194 185
le 2). Finally, data for demersal species from the
Fig. 1. Sampling sites in Augusta Bay and distribution of total mercury (HgT) in bottom sediments (data from Sprovieri et al. (2011)).
Table 1Number of specimens per species caught in the sampling sites.
Mid-water sampling (outside the bay) Bottom-water sampling (inside the bay)
Species C1 (no.) C2 (no.) Habitat Species C3 (no.) C4 (no.) Habitat
Engraulis encrasicolus 20 20 Pelagic Diplodus annularis 59 15 DemersalSardina pilchardus 8 20 Pelagic Diplodus vulgaris – 3 DemersalBoops boops – 20 Pelagic Pagellus erythrinus 1 6 DemersalTrachurus trachurus – 6 Pelagic Pagellus acarne 11 1 DemersalIllex coindetii 6 – Pelagic Sepia officinalis 2 6 DemersalLoligo forbesi 3 – Pelagic Serranus scriba 2 – DemersalPagellus erythrinus 1 – Demersal Caranx rhonchus 1 – PelagicPagellus bogaraveo 2 Demersal Sphyraena sphyraena 1 PelagicMullus barbatus 1 – Benthic Scorpaena notata – 5 Benthic
Scorpaena scrofa 3 1 BenthicMullus barbatus – 3 BenthicMullus surmuletus 1 1 BenthicMurena helena 1 – BenthicOctopus vulgaris – 1 Benthic
186 M. Bonsignore et al. / Food and Chemical Toxicology 56 (2013) 184–194
Table 2HgT means concentrations in muscle and liver of the analysed species and comparison with data for other areas.
Species No. Range of total length(mm)
HgT muscle(lg g�1)
S.D. References Site HgT liver(lg g�1)
S.D.
Engraulisencrasicolus
40 109–138 0.052 0.019 This work Augusta 0.204 0.147
11 120–139 0.057 0.014 This work Marsala 0.119 0.0380.040 Bilandzic et al. (2011) Adriatic sea
9 121–147 0.070 0.090 Gibicar et al. (2009) Adriatic seaa
0.030 0.030 Copat et al. (2012) Sicily (Catania)0.060 0.030 Copat et al. (2012) Syracuse (Sicily)
18 0.060 Pastor et al. (1994) Mediterranean sea(Spain)a
0.070 Martorell et al. (2011) Mediterranean sea(Spain)
4 0.055 0.003 Tuzen (2009) Black sea (Turkey)a
Sardina pilchardus 28 115–150 0.082 0.035 This work Augusta 0.196 0.15710 168–178 0.090 0.040 Gibicar et al. (2009) Adriatic seaa
0.080 0.030 Copat et al. (2012) Catania (Sicily)0.180 Buzina et al. (1995) Adriatic seaa
0.198 Buzina et al. (1995) Adriatic sea (KastelaBay)a
14 0.052 Wolfgang (1983) Adriatic seaa
35 190–260 0.066 Wolfgang (1983) Biscay Bay5 157–165 0.050 Wolfgang (1983) Mediterranean sea
41 0.170 Wolfgang (1983) Ligurian sea28 120–150 0.030 Wolfgang (1983) North Africa (Ceuta)a
20 160–210 0.040 Wolfgang (1983) Western english channel38 0.105 Pastor et al. (1994) Mediterranean sea
(Spain)a
0.019 Martorell et al. (2011) Mediterranean sea(Spain)
7 188–200 0.033 0.016 Harakeh et al. (1985) Lebanon
Boops boops 20 95–150 0.120 0.049 This work Augusta 0.236 0.19111 158–198 0.196 0.204 Hornung et al. (1980) Israela
1 0.075 Pastor et al. (1994) Mediterranean sea(Spain)a
2 130–160 0.190 Stoeppler and Nürnberg(1979)
Med. sea (Dubrovnik)
0.267 Buzina et al. (1995) Adriatic sea0.312 Buzina et al. (1995) Adriatic sea (Kastela
Bay)a
16 139–171 0.036 0.025 Harakeh et al. (1985) Lebanon
Trachurustrachurus
6 56–222 0.131 0.147 This work Augusta 0.344 0.176
2 260–285 0.170 Stoeppler and Nürnberg(1979)
North sea (German Bight)
170 0.170 Mikac et al. (1984) Adriatic sea (KastelaBay)a
37 130–236 0.122 0.101 Hornung et al. (1980) Israela
16 159–203 0.045 0.019 Harakeh et al. (1985) Lebanon0.053 Martorell et al. (2011) Mediterranean sea
(Spain)4 0.078 0.005 Tuzen (2009) Black Sea (Turkey)a
5 0.053 0.012 Keskin et al. (2007) Marmara sea (Turkey)a
Diplodus annularis 74 109–179 0.557 0.303 This work Augusta 1.195 0.8270.653 Buzina et al. (1995) Adriatic seaa
0.628 Buzina et al. (1995) Adriatic sea (KastelaBay)a
Diplodus vulgaris 3 102–179 0.643 0.614 This work Augusta 2.035 2.5545 0.378 0.017 Keskin et al. (2007) Marmara sea (Turkey)a
Sphyraenasphyraena
1 1190 2.269 This work Augusta 9.727
14 219–295 0.167 0.068 Hornung et al. (1980) Israela
Caranx rhonchus 1 264 1.701 This work Augusta
Pagellus acarne 12 149–161 0.254 0.028 This work Augusta 0.618 0.1783 135–141 0.112 Hornung et al. (1980) Israela
15 164–182 0.032 0.014 Harakeh et al. (1985) Lebanon
Pagellus bogaraveo 2 178–179 0.266 0.227 This work Augusta 1.230 0.700
Pagellus erythrinus 8 154–205 0.407 0.100 This work Augusta 2.322 0.4455 110 0.341 0.025 Papetti and Rossi (2009) Tyrrhenian sea (Lazio)
57 115–187 0.180 0.094 Hornung et al. (1980) Israela
9 89–173 0.240 0.190 Gibicar et al. (2009) Adriatic seaa
(continued on next page)
M. Bonsignore et al. / Food and Chemical Toxicology 56 (2013) 184–194 187
infonth2(T
3
insTati(TTTsT
found between the same or similar species, collectedllus acarne, Pagellus erythrinus, M. barbatus, Mullus sur-d ouy.
n
bio
ccuioke co, 20
anratian
ecielly
Table 2 (continued)
Species No. Range of total length(mm)
HgT muscle(lg g�1)
S.D. References Site HgT liver(lg g�1)
S.D.
28 140–152 0.042 0.023 Harakeh et al. (1985) Lebanon5 0.168 Pastor et al. (1994) Mediterranean sea
(Spain)a
5 0.290 0.044 Keskin et al. (2007) Marmara sea (Turkey)a
Serranus scriba 2 122–140 2.165 0.768 This work Augusta 2.581 0.5923 1.030 0.459 Gibicar et al. (2009) Tyrrhenian sea
(Tuscany)a
Mullus barbatus 4 155–202 0.815 0.777 This work Augusta 1.518 0.582102–230 0.116 0.056 Hornung et al. (1980) Israela
0.400 0.400 Storelli et al. (2004) Ionian sea0.490 0.500 Storelli et al. (2004) Adriatic seaa
13 117–180 0.700 0.730 Gibicar et al. (2009) Adriatic seaa
0.370 Buzina et al. (1995) Adriatic sea0.318 Buzina et al. (1995) Adriatic sea (Kastela
Bay)a
59 0.139 Pastor et al. (1994) Mediterranean sea(Spain)
0.010 Martorell et al. (2011) Mediterranean sea(Spain)
30 128–166 0.054 0.025 Harakeh et al. (1985) Lebanon130–200 0.233 Stoeppler and Nürnberg
(1979)Mediterranean Sea(Sardinia)
4 0.036 0.002 Tuzen (2009) Black Sea (Turkey)a
5 0.434 0.012 Keskin et al. (2007) Marmara sea (Turkey)a
Mullus surmuletus 2 200–209 0.662 0.089 This work Augusta 1.1129 120–160 0.086 Hornung et al. (1980) Israela
59 0.139 Pastor et al. (1994) Mediterranean sea(Spain)a
2 185–203 0.250 Stoeppler and Nürnberg(1979)
North sea (German Bight)
37 0.060 Bilandzic et al. (2011) Adr.sea (Croatian coast)
Scorpaena scrofa 4 93–112 1.082 0.285 This work Augusta 1.637 0.3800.222 Buzina et al. (1995) Adriatic sea0.390 Buzina et al. (1995) Adr. sea (Kastela Bay)a
Scorpaena notata 5 114–133 1.340 0.380 This work Augusta 2.339 0.5295 0.490 0.430 Gibicar et al. (2009) Tyrrhenian sea
(Tuscany)a
Illex coindetii 6 33–92 0.078 0.039 This work Augusta13 52–224 0.100 0.100 Gibicar et al. (2009) Adriatic seaa
Loligo forbesi 3 45–170 0.147 0.024 This work Augusta 0.311 0.011
Sepia officinalis 8 108–148 0.766 0.288 This work Augusta
Octopus vulgaris 1 123 0.443 This work Augusta
188 M. Bonsignore et al. / Food and Chemical Toxicology 56 (2013) 184–194
ner bay show the widest non-outlier ranges (0.084–1.116 lg g�1
r muscles, 0.109–2.747 lg g�1 for livers) and the most elevatedumber of outliers and extreme values (Fig. 2a and b). In particular,e highest HgT mean values (2.165 lg g�1 in liver and
.581 lg g�1 in muscle) were measured in the S. scriba speciesable 2).
.3. THQ and EWI values
Mean THQ and EWI values calculated for each caught species,side and outside the bay, are reported in Table 3. Most of the fish
pecies inside the bay show higher values (TQa = 1.53–15.8;Qb = 0.66–6.88; EWI = 1.06–11.0) than those outside the studiedrea (TQa = 0.31–4.20; TQb = 0.66–6.88; EWI = 0.22–2.91). In par-cular, the highest values were calculated for M. helenaQa = 15.8; TQb = 6.88; EWI = 11.0), S. scriba (TQa = 13.0;
Qb = 5.65; EWI = 9.02) and Caranx rhonchus (TQa = 10.2;Qb = 4.44; EWI = 7.09) caught inside the bay, while, the pelagicpecies outside the bay show the lowest values (TQa = 0.31–0.88;
ences wereinside (Pagemuletus) anatus) the ba
4. Discussio
4.1. Mercury
The Hg aimportant bprey, uptakrate (Wang1999; DangHg concentetc.) (Wangchemical spstill not fu
Murena helena 1 800.5 2.638 This work
a Polluted site.
Qb = 0.14–0.38; EWI = 0.22–0.61). Finally, no significant differ- 2002; Wang an
Augusta 3.817
tside (Pagellus bogaraveo, P. erythrinus and M. barb-
accumulation effects: length/age vs. HgT content
mulation in marine fish primarily depends on someinetic parameters: assimilation from the ingestednstants from the aqueous phase, de-toxification12; Wang et al., 1997, 1998; Wang and Fisher,d Wang, 2011) and environmental features (e.g.,on and speciation in seawater, dietary sources,d Wong, 2003). However, physiological and geo-s-specific influences on Hg bioaccumulation are
understood (Baines et al., 2002; Xu and Wang,
d Wong, 2003; Dang and Wang, 2012).than20peseanontr(SleG
ortedynnu001, 20199ions coLuteod.twecrea
elag
M. Bonsignore et al. / Food and Chemical Toxicology 56 (2013) 184–194 189
Several studies have demonstrated that Hg concentrations ine muscles of marine organisms proportionally increase with sized age (Lange et al., 1994; Burger et al., 2001; Green and Knutzen,03; Simonin et al., 2008). Moreover, Hg de-toxification rates ap-ar negatively correlated with the fish size (Trudel and Rasmus-n, 1997), supporting a potential correlation between Hg levelsd size/age in the organisms. However, detailed investigationsdifferent groups of species and on a wide range of HgT concen-
ations are lacking and, when available, sometimes controversialtafford and Haines, 2001), especially for fishes with low mercuryvels (average below 0.2 ppm) (Park and Curtis, 1997; Burger and
fish are rep(Thunnus thcotrigiano, 2Sea (Storelli(Joiris et al.,tive correlatmarine fishePacific) andin Atlantic Crelations beslight Hg in
Fig. 2. Box-plots with HgT concentrations in the muscles and livers of p
ochfeld, 2011). Strong correlations between size and Hg levels in northern Tyrrhe
for Swordfish (Xiphias gladius) and Bluefin Tunas) from the Mediterranean Sea (Storelli and Mar-), for several pelagic fish species from the Adriatic08) and for S. pilchardus specimens from Tunisia9). Furthermore, Burger et al. (2007) found a posi-between size and Hg levels for 11 of 14 species ofllected in the western Aleutians (Bering Sea/Northn et al. (1987) found the same positive correlation
Moreover, Leonzio et al. (1981) report positive cor-en Hg content and weight in M. barbatus and asing trend with size in E. encrasicolus from the
ic, demersal and benthic fishes.
nian Sea. Finally, Gewurtz et al. (2011) show strong
cw
bolemwee2nrs(r(r(rracn
supiedstudatasconandNOilab
of H
arees b
tedd, Hut man
onateeve
Table 3THQ and EWI calculation for each caught species (inside and outside the Bay).
Inside Outside
Species THQa THQb EWI Species THQa THQb EWI
Caranx rhonchus 10.2 4.44 7.09 Engraulis encrasicolus 0.31 0.14 0.22Diplodus annlularis 3.34 1.45 2.32 Sardina pilchardus 0.49 0.21 0.34Diplodus vulgaris 3.86 1.68 2.68 Boops boops 0.72 0.31 0.50Pagellus acarne 1.53 0.66 1.06 Trachurus trachurus 0.79 0.34 0.55Pagellus erythrinus 2.30 1.00 1.60 Illex coindetti 0.47 0.20 0.33Scophaena scrofa 7.17 3.12 4.98 Loligo forbesi 0.88 0.38 0.61Scorphaena notata 8.05 3.50 5.59 Pagellus bogaraveo 1.60 0.69 1.11Mullus barbatus 5.59 2.43 3.88 Pagellus erythrinus 3.41 1.48 2.37Mullus surmuletus 3.97 1.73 2.76 Mullus barbatus 4.20 1.82 2.91Serranus scriba 13.0 5.65 9.02Murena helena 15.8 6.88 11.0Octopus vulgaris 2.66 1.16 1.85Sepia officinalis 4.60 2.00 3.19
a: USEPA’s reference dose (0.1 lg Hg kg bw�1 d�1).b: acceptable daily intake determined by WHO (0.23 lg Hg kg bw�1 d�1).
F crasth
190 M. Bonsignore et al. / Food and Chemical Toxicology 56 (2013) 184–194
orrelation between HgT concentration and length in most fresh-ater fishes from the Canadian Great Lakes and Ontario (Canada).
Here, the high number of specimens available from pelagic,enthic and demersal fish species associated with a wide rangef length/age and HgT variability detected in tissues offer a chal-nging opportunity to explore in more depth the actual bioaccu-ulation processes of Hg in the studied organisms. In particular,e assumed the length of fishes as a reliable parameter for age
stimates (Boening, 2000; Waldron and Kerstan, 2001; Scuddert al., 2009; Panfili et al., 2010; Basilone et al., 2011; Bacha et al.,012) and, thus, reported HgT values vs. length to assess biomag-ification of that contaminant with time. Statistically reliable andobust correlations were found between HgT mean values mea-ured in muscles for size classes and length in S. pilchardus2 = 0.75), E. encrasicolus (r2 = 0.92), Trachurus trachurus2 = 0.96), Diplodus annularis (r2 = 0.98) and P. erythrinus2 = 0.64) (Fig. 3). Specifically, the calculated HgT accumulation
ates for S. pilchardus, E. encrasicolus, T. trachurus, P. erythrinusnd D. annularis are 0.011, 0.012, 0.022, 0.036 and 0.136 lg g�1 -m�1, respectively, in good agreement with data reported by Hor-
definitivelyfor the studfect on the
In our dtween HgTencrasicolus(p < 0.005; Arange of ava
4.2. Sources
Muscleslevels in fishism associa2004). Indeein muscle, biod (Boudoution onlybioaccumulincreasing l
ig. 3. Relationship between HgT concentrations and total body length for Sardina pilchardus, Engraulis ene mean values for each size class.
ung et al. (1980) for P. erythrinus and T. trachurus species. This clearly reflected
ports a significant linear HgT-length relationshipfish species and a species-specific accumulation ef-ied marine organisms.et an evident increasing trend was measured be-tent and age in the two most abundant species, E.
S. pilchardus (Fig. 4) with significant differencesVA test) among age group, although, the restrictedle age classes needs a larger data collection.
gT and fish contamination in Augusta Bay
the most commonly analysed tissues to monitor Hgecause they represent the edible part of the organ-
with human health risk implications (Henry et al.,g accumulates over time more readily in liver thanuscle appears to retain Hg for a much longer per-
d Ribeyre, 1995). Thus, liver may provide informa-short-term exposure to Hg pollution or mayonly when an organism is exposed to constant orls of dietary mercury (Atwell et al., 1998). This is
icolus, T. trachurus, D. annularis and P. erythrinus. Points represent
in the studied dataset, where HgT concentration
mm
BahilocoHlacewcubygare
l T1 an
ffluxntiaic w
Aubleosalar ethe cchemsystet ag e
Fig. 4. Relationship between total mercury concentration (median value of HgT) in fish muscles vs. age in E. encrasicolus and S. pilchardus. Black lines = confidence interval.
om t
M. Bonsignore et al. / Food and Chemical Toxicology 56 (2013) 184–194 191
easured in liver is up to two orders of magnitude higher than inuscles.The HgT content measured in the tissues of fishes from Augusta
y show an increasing trend with habitat depth, specifically, withghest values measured in benthic species with respect to thewest levels detected in pelagic organisms (Table 2). Additionally,ntamination effects show a south-north gradient evident from
gT levels measured on the ubiquitous Pagellus spp. and D. annu-ris specimens (Fig. 5). In particular, the highest HgT mean con-ntrations occur in fishes caught from southern Augusta Bayhere bottom sediments show the highest concentrations of mer-ry (Sprovieri et al., 2011) (Fig. 1). This suggests a key role playedthe highly polluted sediments as sources of Hg to the investi-
ted marine environment. Also, measurements of HgT in seawater
Internationa0.25 nmol L�
fect of Hg ewith a potein the trophspecies fromby comparasludge disp1980; Gibicunderlyingcific biogeothe studied
Sprovieria potential H
Fig. 5. Differences in muscle (M) and liver (L) HgT contents in Pagellus spp. and Diplodus annularis fr
ported from the bottom, mid and surface waters by ENVIRON nean seawater,
eam (2008) with an average concentration ofd range of 0.05–0.37 nmol L�1 show a crucial ef-from sediments of the bay to the water column
l direct impact on the bioaccumulation processeseb. A direct comparison of HgT content in benthicgusta Bay and other Mediterranean areas affectedHg discharges by chlor-alkali plant and sewage
, specifically Tuscany and Israel (Hornung et al.,t al., 2009), show 2–7 times higher values, thus,ombined effects of high pollution levels and spe-ical pathways driving mercury bioavailability in
em (Table 2).l. (2011) and Fantozzi et al. (2013) have argued onxport from Augusta Bay to the Eastern Mediterra-
he northern (C3) and the southern (C4) part of Augusta Bay.
as a result of the measured 3D circulation system.
NA(Din
soewsJapsfo12Se
s(pobsfiddpTtir
4fi
wethrs(PLintis2Sthclethra(SDa1thcdtubw2o
FAOinlcuin
sideingumpd bis
l ac
ons
n c
HgBaylutef biotamheas
erscfromande A
f themitiv
Inte
ors
gem
archSaluinceInstg. Tola)) foees a
son, Kan
an. J.ellinRegoical dali, Aips bngra
ian csher,raceimnonannida,
andisheto T
nca,nca,nei n
1 (20
onetheless, Hg contamination detected in sediments outside ofugusta Bay, by effects of dredged material from the inner bayi Leonardo et al., 2008; Tranchida et al., 2010), could also directlyfluence the state of pollution of the open sea.
Here, we extend the potential role of Augusta Bay as an Hg pointource for the open ocean also considering the significant transferf pollutants by pelagic fishes moving between the inner andxternal part of the bay. This implies potential effects on the foodeb of the surrounding area as already reported by several authors
tudying marine systems (Riisgard and Hansen, 1990; Futter, 1994;rman et al., 1996; Atwell et al., 1998). This hypothesis is sup-orted by the high mean HgT concentrations measured in pelagicpecies caught outside the bay, which are similar to those reportedr other sites affected by Hg pollution: the Adriatic Sea (Wolfgang,
983; Storelli and Marcotrigiano, 2001; Storelli et al., 2002, 2004,007, 2010; Gibicar et al., 2009), Turkish areas (Tuzen, 2009),panish coastal areas (Pastor et al., 1994) and Israel area (Hornungt al., 1980) (Table 2).
The HgT mean concentrations measured in the livers of E. encra-icolus specimens from Augusta Bay are about twice as high
= 0.044) as those measured in fishes from the unpolluted areaf Marsala (Table 2), suggesting a direct, short-term effect of theay pollution on the pelagic fishes. Moreover, the caught pelagicpecies prefer to inhabit warmer coastal seawaters during theirrst life stages (Basilone et al., 2011), but they generally move ineeper waters during the older stages (Wirszubski, 1953; Schnei-er, 1990; Whitehead, 1990), thus, representing a significant andotential vehicle of contaminants to the deep marine food web.his evidence definitively corroborates our hypothesis of a poten-al Hg export through the food web, from Augusta Bay to the sur-
ounding area.
.3. Target hazard quotient and weekly intake: a real health risk fromshery in the Augusta Bay?
Although estimation of the target hazard quotient (THQ) andeekly intake (EWI), do not provide a quantitative and definitive
stimate on the dangerous health effects on exposed populations,ese methodologies offer preliminary information on the health
isk level resulting from pollutant exposure. Several authorshowed that selenium (Se) offers protection against Hg toxicityarízek and Ostádalová, 1967; Satoh et al., 1985; Ralston, 2009;
émire et al., 2010) that suggests to take in account Se contentsfishes to assess a real risk associated to Hg intake. Positive rela-
onships has been found between Hg and Se contents in differenteawater fishes (Burger and Gochfeld, 2011; Dang and Wang,011; Calatayud et al., 2012). Ralston et al. (2008) showed thate:Hg molar ratios above 1 protect against Hg toxicity. Howeveris ratio definitively depends o species-specific toxic-kinetics pro-
esses (Watanabe, 2002; Burger and Gochfeld, 2012). This featureads to a wide variability of Se:Hg molar ratios and makes difficulteir use in risk assessment. Accordingly, here we estimated health
isk for Hg intake only on THQ and EWI parameters. These indexesre widely used to assess risk associated with fish consumptiontorelli et al., 2004, 2010; Storelli, 2008; Martorell et al., 2011;omingo et al., 2012). In particular, for no carcinogenic effects,n HQ exceeding 1.0 indicates a potential health risk (US EPA,989). In our dataset, either using USEPA’s reference dose (TQa)at WHO acceptable daily intake (TQb), species inside the Bay ex-
eeded the value 1 in all cases, while fishes outside the Bay only inemersal and benthic fishes (P. bogaraveo, P. erythrinus, M. barba-s) (Table 3). International agencies indicate a provisional tolera-
le weekly intake (PTWI) of Hg, ranging from 0.7 lg kg�1 bodyeight (b.w.) (US-EPA, 2004) to 1.6 lg kg�1 b.w. (FAO/WHO,
006). These limits represent safe values for human population
EPA, 2004;the Bay andmary, the caof fish fromhealth of recurrent fishin the consdemersal anin this areatimely socia
5. Conclusi
The maifollows:
� The highAugustafrom poleffects o
� High conzone of tronmentand undtransfer
� The THQinside thhealth osuming dBay defin
Conflict of
The auth
Acknowled
This reseorato dellato express sprophylaticfish samplinCapo GranitCNR, Naplesologies. Thrcontribution
References
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92 M. Bonsignore et al. / Food and Chemical Toxicology 56
ver lifetime. The calculated EWI index exceed the PTWI (US- 30 (1), 19–26.
/WHO) in almost all the species collected insidedemersal and benthic fishes from outside. In sum-lated THQ and EWI highlight that the consumptionside the Bay represents a serious risk for humannt populations and confirm the importance of theban in this area. Also, the results suggest cautiontion of fishes from outside the Bay, especially ofenthic species, confirming that Hg contaminationa serious concern that calls for appropriate andtions.
onclusions of this work can be synthesised as
T concentrations measured in benthic species fromsuggest an active release mechanism of mercury
d sediments to the water column, with consequentaccumulation in the trophic web.ination of pelagic species measured in the external
bay confirms the role of the Augusta marine envi-a potential Hg source for the surrounding area
ores the crucial risk associated with contaminantthe semi-enclosed basin to the open sea.EWI values advise that consumption of fish from
ugusta Bay represents a serious risk for humane local populations, while suggest caution in con-ersal and benthic fishes from outside the Augustaely demanding for appropriate social actions.
rest
declare that there are no conflicts of interest.
ents
is part of an Italian project funded by the ‘‘Assess-te della Regione Siciliana’’. The authors would likere thanks to Dr. M.M. Uccello and Dr. G. Baffo (Zoo-itute of Augusta) for their facilities and support inhanks are also due to Dr. F. Bulfamante (IAMC-CNR,for logistic contribution and Dr. M. Barra (IAMC-
r comments and suggestions on statistical method-anonymous reviewers are warmly tanks for theirnd suggestions.
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Chemosphere 93 (2013) 2024–2032
Contents lists available at SciVerse ScienceDirect
Chemosphere
boundae marinece of po
tmospheric mercury concentrationsaseous elemental mercuryea–air evasionarine boundary layer
ugusta basin
b s t r
e first atsin (SE Sdustrialurce of prmed in tntrationsspectivel
(range 1.1regions els3.6 ± 0.3 (uments to thabout 9.7 ± 0.1 g d (�0.004 t yr ), accounting for �0.0002% of the global Hg oceanic evasion (2000t yr�1). The new proposed data set offers a unique and original study on the potential outflow of Hg from
nd Pacyna, 1998; Kim et al., 2005; Travnikov, 2005; Amos et al.,012). Although a part of the Hg emitted naturally comes from
nd ged fdep). H
issdepo. Inont
sources of Hg tbeen strongly rThe exchange orepresents an ienvironmental turnover of this element (Mason et al., 1994; Lam-
045-6535/$ - see front matter � 2013 Elsevier Ltd. All rights reserved.ttp://dx.doi.org/10.1016/j.chemosphere.2013.07.025
⇑ Corresponding author. Tel.: +39 3471258337; fax: +39 091 6169908.E-mail address: [email protected] (E. Bagnato).
he sea–air exchange of mercury (Hg) in the marine boundary layerf the Augusta basin (southern Italy): Concentrations and evasionux
. Bagnato a,⇑, M. Sproveri b, M. Barra c, M. Bitetto a, M. Bonsignore b, S. Calabrese a, V. Di Stefano b,
. Oliveri b, F. Parello a, S. Mazzola c
DiSTeM, University of Palermo, Via Archirafi 36, 90123 Palermo, ItalyIAMC-CNR, Capo Granitola, Via del Mare 3, 91021 Torretta Granitola, Trapani, Italy
journal homepage: www.elsevier .com/locate /chemosphere
IAMC-CNR, Naples, Calata Porta di Massa, Naples, Italy
i g h l i g h t s
The Hg evasion flux in the Augusta basin marineThe human activity has influenced in the past thThe release of Hg from the Augusta Bay is a sour
r t i c l e i n f o
rticle history:eceived 29 April 2013eceived in revised form 3 July 2013ccepted 8 July 2013vailable online 6 August 2013
eywords:ercury evasion flux
a
Thbainsofocere
ry layer was examined.Hg cycle in the Augusta Bay.
llution for the Mediterranean.
a c t
tempt to systematically investigate the atmospheric mercury (Hg) in the MBL of the Augustaicily, Italy) has been undertaken. In the past the basin was the receptor for Hg from an intenseactivity which contaminated the bottom sediments of the Bay, making this area a potentialollution for the surrounding Mediterranean. Three oceanographic cruises have been thus per-he basin during the winter and summer 2011/2012, where we estimated averaged Hgatm con-of about 1.5 ± 0.4 (range 0.9–3.1) and 2.1 ± 0.98 (range 1.1–3.1) ng m�3 for the two seasons,
y. These data are somewhat higher than the background Hgatm value measured over the land± 0.3 ng m�3) at downtown Augusta, while are similar to those detected in other pollutedewhere. Hg evasion fluxes estimated at the sea/air interface over the Bay range fromnpolluted site) to 72 ± 0.1 (polluted site of the basin) ng m�2 h�1. By extending these measure-e entire area of the Augusta basin (�23.5 km2), we calculated a total sea–air Hg evasion flux of
�1 �1
the sea–air interface at the basin, and it represents an important step for a better comprehension of theprocesses occurring in the marine biogeochemical cycle of this element.
. Introduction
Mercury (Hg) is a chronic pollutant of global concern known toe transported long distances in the atmosphere into remote eco-ystems (Schroeder and Munthe, 1998). Hg flux into the atmo-phere from natural and anthropogenic sources has beeneviewed recently and new estimates for the worldwide distribu-on of anthropogenic emissions have been published (Pacyna
geological aously emittquently re-et al., 2005of yearly empreviouslygenic originsediments c
� 2013 Elsevier Ltd. All rights reserved.
eothermal sources, much of it is recycled Hg previ-rom primary or anthropogenic sources, and subse-osited to terrestrial and ocean surfaces (Ericksenence, as a consequence, a large part of the 2000 tions from natural sources is actually reemission ofsited mercury, much of which has an anthropo-
some instances, it has been discovered that marineaminated by industrial effluents may be secondaryo aquatic ecosystems even though discharge haseduced or has even ceased (Bothner et al., 1980).f Hg between oceanic surfaces and the atmospheremportant process for the atmospheric cycling and
boaqwAcbyglge3020art ythhiabthtecebefathshal(FbytrrethnaoffinAufacuthtoGot(idithatto
searr 20
logy
tion
staeas
basily (
adeCiafe p
s a s, 200ern�1)1), wtaini1 (moutflow of total Hg (HgT) from the Bay to the Augusta
ersyr�1
–2%(12
ns iole ints a
eric
menperearc
Fibachde
4–20
rg et al., 1999). The ability of GEM to reemit from terrestrial anduatic surfaces keeps this element circulating in the environment,ith burial into ocean sediments as the only long term sink.cording to the global model of Hg biogeochemistry proposedMason et al. (1994), the ocean releases about 1/3 of the total
obal Hg emissions to the atmosphere (about 30% of the total bud-t of atmospheric mercury on a global scale) and receives about–70% of the global atmospheric deposition (Lamborg et al.,02). Re-emissions from the ocean of previously deposited Hge dominated by gaseous elemental mercury (Hg0 or GEM � 2000r�1; Mason et al., 1994), the most volatile and long-lived form ofis metal. Its low solubility and high Henry’s Law constant inducegh GEM evasion fluxes from fresh water systems, accounting forout 7–95% of the total estimated atmospheric Hg deposition inat system. By these considerations, currently there is a clear in-nt to increase both qualitative and quantitative knowledge con-rning the processes occurring during the exchange of Hgtween sediments, the overlying water column and sea-air inter-
ce. Within the frame of the IAMC-CNR/ASP program founded bye Regional Health Department of Syracuse, we performed aort-term (1 yr) monitoring study on Hg distribution and evasion-flux in the atmospheric compartment of the Augusta basin
ig. 1), a site of the Mediterranean Sea strongly affected in the pastthe uncontrolled discharge of Hg (since the 1950s) from indus-
ial and petrochemical plants (Sprovieri et al., 2011). This workpresents an important step toward a better comprehension ofe processes occurring once Hg is re-emitted from the contami-ted bottom sediments (which actually represent the main sourceHg for the Bay; Sprovieri et al., 2011) in the water column andally to the atmosphere. Although the water surface area of thegusta basin represents only a small part of the total oceanic sur-
ces on Earth, we aimed this study may improve the global mer-ry budget and cycle which lack measurements in large parts ofe world’s marine environments. Our purposes are threefold: (1)
characterize the regional background level of atmosphericEM as well as evasional fluxes of Hg in the Bay and compare with
intensive reand summe
2. Methodo
2.1. Site loca
The Augu30 km of thepart of theplants of Itaemission mDonne andcharge of thwhich act aBay (ICRAMin the south527.3 mg kg23.8 mg kg�
ments con12.7 mg kg�
An averagedcoastal wat0.162 kmolsponds to 1nean areaconsideratioimportant rand represe
2.2. Atmosph
Measure(GEM) wereian CNR res
E. Bagnato et al. / Chemosphere 93 (2013) 202
her areas at various latitudes; (2) to evaluate the regional sourcesf any) eventually affecting the GEMatm concentrations; and (3) toscuss the deposition of atmospheric Hg in the Augusta area. Withese aims, a dynamic flux chamber coupled with a real-timeomic adsorption spectrometer (Lumex-RA 915+) has been usedmeasure Hg evasion flux at the sea/air interface during three
graphic cruises(November 201chosen out of prplicity, GEM, Hgcle, unless othewas performed
g. 1. Maps showing (a) the Augusta basin and the site of the installed weather station used in this study,sin, (c) the land trajectories to detect GEM contents in the atmosphere along the coastal area, and (d) thamber technique (ST1–7). Map (c) also reports the partition of the entire basin into seven Voronoi polygtail).
ch oceanographic cruises performed in the winter11–2012.
basin is within a natural Bay which occupies abouttern coast of the Sicily (Fig. 1a). The southernmostn hosted one of the most important chlor-alkaliSyndial Priolo Gargallo), which with 765 kg of Hgup over 20% of total Italian emissions in 2001 (Leani, 2008). The effects of indiscriminate Hg dis-ast include high Hg levels in bottom sedimentsource of Hg to the water column in the Augusta5; Sprovieri et al., 2011). In detail, sediments lyingpart of the Bay show the highest HgT levels (0.1–with a large range of variability (median valuehile the northern area is characterized by sedi-
ng HgT concentrations varying from 0.1 toedian value 1.1 mg kg�1) (Sprovieri et al., 2011).
32 2025
has been estimated to be in the order of about(�0.032 t yr�1; Sprovieri et al., 2011), which corre-of the amount calculated for the entire Mediterra-.5 kmol yr�1; Rajar et al., 2007). By theset emerges that the Augusta basin may play ann exporting mercury into the Mediterranean Seapoint source for that system.
GEM measurements
ts of atmospheric elemental gaseous mercuryformed across the Augusta basin on-board the Ital-h vessel Luigi Sanzo, during three main oceano-carried out along the same route in the winter
1) and summer (July 2011–June 2012) (Fig. 1b),actical and logistical criteria. For the sake of sim-and Hg0 are used without distinction in this arti-rwise specified. The analysis of atmospheric GEMusing an automated real-time atomic adsorption
(b) the routes of the oceanographic cruises inside the Augustae stations for Hg evasion flux measurement by accumulation
ons, each relative to one station of measurement (see text for
s2dwaTp1(Gtrmb�rssestic�srbthcG5otyee2mafr
2
b
partc ada–aberof e
entandPleele
espeaticha
gesghtogenspe
oninwen wx ceaav
t peivelorsse coseo th
o �
is t); (CCo)are
2 24–2
pectrometer (Lumex-RA 915+). The Lumex sampled air at about0 l min�1 directly into the instrument’s sampling inlet (�3 cmiameter) at ambient temperature to the multi path detection cellhich has an effective path length of 10 m. This multi-path cell hasvolume of 0.7 l and air changes in the cell 3 times every 7–10 s.
he instrument inlet has an external washable dust filter with aorosity of 5–10 mkm in addition to a coarse dust filter of porosity00 mkm. The Lumex monitored gaseous elemental mercury
EM) concentrations using differential atomic absorption spec-ometry with correction for background absorption via the Zee-an Effect (Sholupov et al., 2004). A zero correction resets the
aseline every 5 min during sampling. The detection limit was2 ng m�3, and the instrument has an accuracy of 20%. The accu-
acy and precision of the applied instrumentation has also been as-essed through comparison with the traditional gold trap/CVAFSystem used at remote sites elsewhere (Aiuppa et al., 2007; Wittt al., 2008). During the cruises, the air inlet of the analyzer was in-talled on the upper deck about 3 m a.s.l. to avoid the contamina-on from ship emissions. We sampled air at 1 s intervals byovering a total marine area of about 60 km at a speed of10 km h�1. Lumex has also been employed to measure atmo-
pheric GEM concentration over the land along the shoreline sur-ounding the Augusta basin (Fig. 1c), in order to assessackground Hg levels in the air masses inland. For this purpose,e inlet of the instrument unit was connected to a 1 m-long sili-
one tube and mounted outside of a side window of the vehicle.EM concentrations in air were thus continuously quantified ats intervals, by covering about 20 km of the coast by car at a speed
f �20 km h�1. This analyzer has been successfully used in variouspes of atmospheric mercury measurement campaign (e.g., Kim
t al., 2006; Špiric and Mashyanov, 2000; Engle et al., 2006; Wangt al., 2006; Aiuppa et al., 2007; Muresan et al., 2007; Witt et al.,008; Ci et al., 2011) where it was successfully used for continuouseasurements of Hg distribution in the atmosphere over large city
reas and various geological contexts both by walking traverse and
submergedtime atomimate the selation chamlaboratoriesto the differ1995; Carpi2006). Theand UV wavand UV-B, rfor the formThe floatingwith the edensure a tiair. To hominstalled, suAfter positicontact, weconcentratioduce the fluters of the sspeed and wonly a shorduring relatof these factreduces noisurface expaccording t2001):
UGEM ¼ QðC
where UGEM
(ng m�2 s�1
air exiting (is the basal
026 E. Bagnato et al. / Chemosphere 93 (2013) 20
om moving vehicles.
.3. Mercury flux assessment at the sea–air interface
For the first time in this area, we used a plexiglass open-ottom dynamic flux chamber (emerged part: 50 � 50 � 50 cm;
air flowing throtration differentthe system blanmeasured in theclean surface (aprotocol of the
Fig. 2. Positioning (a-b), testing (c) and in real-time measurements (d) of sea–air GEM evasion
: 50 � 50 � 30 cm) technique coupled with a real-sorption spectrometer (Lumex-RA 915+) to esti-
ir Hg evasion flux in the MBL (Fig. 2). The accumu-was built by the technical staff working in thelectronic at IAMC-CNR (Capo Granitola), accordingschemes proposed in literature (Kim and Lindberg,
Lindberg, 1998; Covelli et al., 1999; Wang et al.,xiglas was selected since it transmits all visiblengths in solar radiation (89% and the 64% of UV-Actively; Wang et al., 2006), which are responsibleon of photo-induced gaseous mercury in water.mber system was placed on the sea water surfaceof the chamber immersed 30 cm into the water toseal with the water, preventing entry of outsideize the air inside the chamber two fans have been
nded at about 5 cm from the top of the chamber.g the chamber on the surface and achieving good
re able to reach a steady-state of internal mercuryithin approximately 10 min. This allowed us to re-
hamber’s influence on the environmental parame-water surface we were investigating, mainly windes, because the chamber remains on the water forriod of time. Of course, this technique is suitabley calm conditions of the sea, when the influenceis negligible; anyway, the large size of the chamberaused by the waves. Mercury flux from the waterd in the chamber (0.25 m2) was then calculatede Eq. (1) (Lindberg and Price, 1999; Zhang et al.,
CiÞ=A ð1Þ
he GEM total emission rate per area and unit timeo � Ci) is the difference in GEM concentrations inand entering (Ci) the chamber (DC) (in ng m�3); Aa of the chamber in m2; and Q is the flow rate of
3 �1
032
ugh the chamber in m s . Of course, the concen-ial used in the flux calculation must be greater thank, which we determined based on the DC difference
sunlight by sealing the chamber bottom to a largeclear polycarbonate plate in our case). The QA/QCexperiment has been achieved in the field using
flux by using the accumulation chamber technique.
blwer(Gre19thcerealacat
2.
(wIvit(FfuSa19prabthboSadrMsawintrbuhecatiis
3.
3.
atthPrficpestrotireizNidtith
3.
ha20
nd s�C.
c(ran(Tato steraierillectnatu
keycapeinte. Timthad a
m�3
theasinme
EMce gatay,
enttedal., 2
ilaay,d 2al.,
e prrran
4–20
anks. Before and after each oceanographic cruise, the chamberas extensively cleaned with diluted laboratory detergent and sev-al-fold rinsed with Milli-Q water. We find negligible blank valueEM blank � 0.15 ng m�3) which agrees well with the blank testsported in literature (�0.2 ng GEM m�3; Carpi and Lindberg,98; Gustin et al., 1999). The theoretical Hg concentration (frome manual calibration) has been compared to the measured con-ntration by direct injection into the analytical device and thecovery rate from direct injection into the flux chamber. The over-l QA/QC protocol showed that up to 99% of the accuracy has beenhieved by the technical protocol, as also confirmed by the low rel-ive standard deviation exhibited by our data.
4. Bulk deposition collection
Bulk deposition was collected using a glass-made open collectoret + dry deposition) according to Iverfeldt (1991) and Jensen and
erfeldt (1994), which was located on the roof of the port author-ies office close to downtown Augusta close to the weather stationig. 1a). Rainwater samples were collected at irregular intervals innction of the rainy events which affected the examined area.mples were analyzed for total mercury concentrations (OSPAR,97) by a direct analyzer (Milestone DMA-80), which uses theinciple of thermal decomposition, amalgamation and atomicsorption, in operation at IAMC-CNR (Capo Granitola). Beforee analysis, each rainwater sample was weighed into a quartzat, and transferred from the analytical balance to the DMA-80.mple boats, loaded onto the instrument auto-sampler, are firstied and then thermally decomposed in a oxygen-rich furnace.ercury and other combustion products are released from themple and they are carried to the catalyst section of the furnace,here nitrogen and sulfur oxides, as well as halogens and otherterfering compounds, are eliminated. Mercury is selectively
vey, the wiabout 12–25spheric GEM�2.1 ± 0.98respectivelytions showported in li2003; Sprovity in our cosity of therepresents aand then esanyway thein this studywhat higherover the lan1.1 ± 0.3 ngtected alonging the bsporadicallyEstimated Gtoxicity sinbeen fixed2002). Anywto some extvalues reporLindberg etwhile are simlike Tokyo B2.8 ± 1.5 an2006; Fu etin the rangthe Medite
E. Bagnato et al. / Chemosphere 93 (2013) 202
apped, in a separate furnace, through gold amalgamation. Com- 2010) and the Atlanthe
, whc Ocinallruiseheed:latitl gase/oc
ange1999zziericatmomposphng1). S
f thes G
GEM
leavpere ahasthe
grap
stion by-products are flushed off. The amalgamation furnace isated and mercury is rapidly released. Mercury is flown via therrier gas into a unique block with a tri-cell arrangement, posi-
oned along the optical path of the spectrophotometer, where itquantitatively measured by atomic absorption at 253.65 nm.
Results and discussions
1. Meterological pattern of the area
A meteorological data set, including wind direction, air temper-ure, and precipitation amount, was developed using data frome continuous acquisition by a weather station (DAVIS – Vantageo 2 Wi-Fi) installed on the roof of the Augusta port authorities of-e (Fig. 1a). Wind rose diagrams, median wind direction and dis-rsion parameters were computed by means of ‘‘openair’’ R
atistical package. The overall (whole observation period) windse diagram (Fig. 3a) showed that the most frequent wind direc-
ons were related to NW and NE sectors. Winds from NW sectorpresents about the 50% of total observations, and were character-ed by prevalent N-W direction (25%), while the prevalent one forE sector (accounting for 35% of the total) was E-NE. We did notentify any relevant relation between the measured Hg concentra-ons in the MBL and meteorological parameters recorded duringe survey.
2. GEM distribution in the MBL
The GEM measurements in the MBL over the Augusta basinve been performed along the same route in winter (November
2008), the A2003) andet al., 2003)torial Pacifi(Table 1). FOCEANOR c(2013) in tm�3, averagfunction ofspheric totaother marinSlemr and Lborg et al.,2011; Fantoter-hemisphof Hg to thesphere. In cnormal atm95–98% amoCi et al., 201face, most oconsidered a
3.3. Air–sea
Mercurypresent at surespect to thevasion fluxlected alongthe oceano
11) and summer (July 2011–June 2012) (Fig. 1b). During the sur- therefore the ef
peed ranged from 4.5 to 9.8 m s�1 and Tair fromDuring the cruises we measured averaged atmo-
oncentrations of �1.5 ± 0.4 (range 0.9–3.1) andge 1.1–3.1) ng m�3 in the winter and summer,
ble 1). The time-weighted average GEM concentra-ome extent seasonal variations, as previously re-ture for other geographic areas (Sprovieri et al.,and Pirrone, 2008; Wangberg et al., 2008). Variabil-ed Hg data may be ascribed to the different inten-ral sunlight between winter and summer whichparameter in controlling rates of % Hg0 producedd from seawater surface (Costa and Liss, 1999);nsity of solar radiation has not been determinede series evidence that collected data result some-
n the atmospheric background Hg level measuredt the downtown urban site of Augusta (averaged; Table 1 and Fig. 4a), while are similar to those de-shore close to the dense industrial area surround-
(range 1.5 ± 1.4–2 ± 1.6 ng m�3), where weasured Hg peaks of about 8–10 ng m�3 (Fig. 4a).at background levels suggests no significant acuteenerally the lowest adverse effect observed has15–30 lg m�3 (�15–30 � 103 ng m�3; Kazantzis,GEMatm level over the Augusta basin results are
higher than the background atmospheric mercuryfor the North Hemisphere (range: 1.5–1.7 ng m�3;007; evidenced by the dashed red lines in Fig. 4b),
r to those reported for a few polluted marine areas,the South China Sea and the Yellow Sea (1.9 ± 0.6,.3 ± 0.7 ng m�3, respectively; Narukawa et al.,2010; Ci et al., 2011; Table 1). Our data are alsooposed for many other oceans and seas, such asean (1.5 ± 0.3–2 ± 0.6 ng m�3, Sprovieri et al.,driatic Sea (1.6 ± 0.4 ng m�3, Sprovieri and Pirrone,tic Ocean (1.3 ± 0.1–2 ± 0.1 ng m�3, Temme et al.,North Pacific Ocean (2.5 ± 0.5 ng m�3, Laurier
ile are higher than those measured over the equa-ean (1.0 ± 0.1 ng m�3, Kim and Fitzgerald, 1986)y, our data fit with results from the 2010 MED-
campaign recently performed by Fantozzi et al.Eastern Mediterranean (range: 1.3–1.8 ng GEM1.6 ± 0.1 ng m�3; Table 1). By displaying data as aude (Fig. 5) and for comparison with the atmo-eous mercury (TGM = GEM + RGM) contents fromeanic environments (Kim and Fitzgerald, 1986;r, 1992; Fitzgerald, 1995; Mason et al., 1998; Lam-; Narukawa et al., 2006; Fu et al., 2010; Ci et al.,
et al., 2013) we found a small but discernible in-gradient in GEM resulting from greater emissions
osphere in the more industrialized Northern Hemi-iling this diagram we considered that under theeric condition, GEM is generally taken more than
all the atmospheric Hg species (i.e. RGM and Hgp;ince RGM is easily adsorbed by the seawater sur-literature reported TGM measurements should be
EM.
flux
es the ocean by evasion of dissolved Hg0 when it issaturated concentration in the surface waters withtmosphere (Kim and Fitzgerald, 1986). Sea–air Hg
been measured at seven monitoring stations se-Augusta basin (ST1-7; Fig. 1c and Table 2). Duringhic cruises, weather conditions were optimal,
32 2027
fect of wind, waves and presence of clouds were
nfrpH(stwflcinatiomhsF
: 0.011trib
entourts (oned f198al.,m�2
m�
t al), ththe
Fig. 3. Wind rose diagram built for the whole observation period showing the most frequent wind directions. Winds from NW sector represents about the 50% (c) of totalobservations, and were characterized by prevalent N-W direction (25%) (b), while the prevalent one for NE sector (accounting for 35% of the total) was E-NE (a).
Table 1Gaseous elemental mercury (GEM) concentrations measured in the MBL over the Augusta basin compared to literature data for other acquatic environments. For a more detaileddescription on averages and methods the reader is referred to the original article.
Measurement sites Period GEM (ng m�3) S.D. (n) Methods References
Range Mean
Augusta Basin MBLWinter period averaged 2011/11/29–30 (0.9–3.1) 1.5 0.4 (8137) Lumex RA-915 + analyzer Present studySummer period averaged 2011/07/11–12–2012/06/23–24 (1.1–3.1) 2.1 0.98 (159) Lumex RA-915 + analyzer Present studyAugusta downtown 11/11/2011 (0.8–1.4) 1.1 0.3 (129) Lumex RA-915 + analyzer Present study
Other sitesSweden coastal areas October 1979–September 1980 (2.7–4) 3.4 0.4 (12) Gold-traps Brosset (1992)North Atlantic ocean October 1977–January 2000 (1.7–2) 2 0.1 (8) Tekran 2537A analyzer Temme et al. (2003)South Atlantic ocean October 1977–February 2001 (1–1.5) 1.3 0.1 (10) Tekran 2537A analyzer Temme et al. (2003)South Atlantic 1996/05/20–1996/06/17 (1.2–1.9) 1.6 0.2 (14) Gold-traps Lamborg et al. (1999)North pacific ocean 2002 (1.6–4.7) 2.5 0.5 (n.a.) Tekran 2537A analyzer Laurier et al. (2003)Equatorial pacific ocean 1984/07/03–1984/06/08 (0.8–1.1) 1.0 0.008 (23) Gold-traps Kim and Fitzgerald (1986)Indian ocean 2007 (1–1.5) 1.2 0.06 (n.a.) Tekran 2537A analyzer Witt et al. (2010)Eastern Mediterranean sea August 2003–September 2006 (1.3–2) 1.5 0.3 (5) Tekran 2537A analyzer Sprovieri et al. (2010)Western Mediterranean August 2003–July 2007 (1.2–2.7) 2 0.6 (3) Tekran 2537A analyzer Sprovieri et al. (2010)East Mediterranean 2010/08/26–2010/09/13 (1.3–1.8) 1.6 0.1 (15) Tekran 2537A analyzer Fantozzi et al. (2013)Baltic sea 1997/07/02–15 (1.4–2) 1.7 0.2 (11) Tekran 2537A analyzer Wängberg et al. (2001)Baltic sea 1998/03/02–15 (1.2–1.6) 1.4 0.1 (9) Tekran 2537A analyzer Wängberg et al. (2001)Adriatic sea 2004/10/26–2004/11/12 (0.8–3-3) 1.6 0.4 (n.a.) Tekran 2537A analyzer Sprovieri and Pirrone (2008)Tokyo Bay 2003/12, 2004/10, 2005/01 (1.1–2.8) 1.9 0.6 (22) Automated Hg analyzer Narukawa et al. (2006)South China sea 2008/05/09–2009/05/18 (1.5–4.5) 2.8 1.5 (n.a.) Tekran 2537A analyzer Fu et al. (2010)
6)
n
2028 E. Bagnato et al. / Chemosphere 93 (2013) 2024–2032
ot taken in consideration in discussing results. Our data rangeom 3.6 ± 0.3 (unpolluted site) to 72 ± 0.1 ng Hg m�2 h�1 (mostolluted site) (Table 2), indicating that the sea–air evasion flux ofg from the basin is not uniformly distributed but varies spatiallyee Fig. 1c and Appendix I), while any particular trend across theo seasons (November 2011–June 2012) has been observed. Each
ux value is devoid of the blank effect, since we subtracted thehamber blank value. The higher Hg evasion fluxes were estimated
the southern part of the basin (sampling stations ST1 and ST3,ccounting for about 36 ± 0.3 and 72 ± 0.1 ng Hg m�2 h�1, respec-vely; Fig. 1c), the most contaminated area of the basin in termsf Hg contained in the bottom sediments (0.1–527.3 mg Hg kg�1,edian value 23.8 mg Hg kg�1; Sprovieri et al., 2011). On the other
and, the lowest Hg flux has been measured close to the northernector of the basin (ST4, 3.6 ± 0.3 ng Hg m�2 h�1; Table 2 and
tents (rangeieri et al., 2are key conmay represcomparingenvironmensulted to beues reporteFitzgerald,Fantozzi et2.4 ± 1.5 ng(4.2 ± 3.2 ngAndersson eet al., 20082010), and
Yellow sea July 2007–May 2009 (1.12–7) 2.3 0.7 (120
.a. = Not available.
ig. 1c), where the bottom sediments exhibited quite low Hg con- 2006). In detai
1–12.7 mg kg�1; median value 1.1 mg kg�1; Sprov-). These results suggest that the marine sedimentsutors of Hg to the marine ecosystem and hencea potential source of Hg to the atmosphere. By
data with literature cases for many marineTable 2), GEMatm flux over the Augusta basin re-
order of magnitude higher than the averaged val-or the Pacific Ocean (3 ± 2 ng m�2 h�1; Kim and6), the Mediterranean Sea (2.2 ± 1.5 ng m�2 h�1,2013; 2.5 ± 1.2 ng m�2 h�1, Gardfeldt et al., 2003;h�1, Ferrara et al., 2000), the Tyrrenian Sea
2 h�1, Gardfeldt et al., 2003; 1.6 ± 1.3 ng m�2 h�1,., 2007), the Artic Ocean (2.4 ng m�2 h�1, Anderssone South China Sea (4.5 ± 3.4 ng m�2 h�1, Fu et al.,Tokyo Bay (5.8 ± 5 ng m�2 h�1; Narukawa et al.,
Lumex RA-915 + analyzer Ci et al. (2011)
l, our results are comparable both to the Hg flux
value estimated2011) and the1998), this last ccury levels in wlisted worldwidindirectly by usi1974); while onvalues estimateThe use of thereal-time atomirepresents an imod to assess Hg2006). This techgoodness of prostrongly dependtions (Wanninkhand Slater, 1974ferent sources wcontrolling emisstrates with a wframework for sAs listed in Tablfrom acquatic enbackground unccomparable to tsubstrates assocHg m�2 h�1, respin areas of thermdiffuse Hg-bearvated Hg conceand Buseck, 198face in the Augulower than Hg rdeposits and meative to naturalaveraged fluxesto tens of thousTable 2). In ordethe entire surfacused the modelmer (1991) (theus to split the baccounting for aestimated a cumabout 0.004 t�0.0002% of theposed by Masonresults to be lowTokyo Bay (ranet al., 2006), buthe water surfac% of the total ocSharman, 2010)
3.4. Bulk deposit
Long-range twet and dry desupplied to terreOur preliminary(from August 20II) are comparabNorth Pacific ONorth Sea (30 ngcorrelation witha chlorine causti
Fig. 4. Time series atmospheric GEM concentration measured (a) in the atmosphereover remote, industrial and urban areas close to the Augusta basin, and (b) at theMBL above the basin. (a) The atmospheric background Hg level measured over theland at the downtown urban site of Augusta is quite low (averaged0.9 ± 0.5 ng m�3), while we measured GEM concentrations peaks of about 8–10 ng m�3 along the coastline close to the dense industrial area surrounding thebasin. (b) The yellow and grey areas indicate the concentration range of GEMmeasured over the Tokyo Bay (range: 1.3–2.5 ng m�3) and the South China Sea(range: 2.1–3.1 ng m�3), respectively, compiled by literature data (Narukawa et al.,2006; Fu et al., 2010). Blue dashed line indicates the averaged GEM value reportedfor the atmosphere over the polluted area of the Yellow Sea (Ci et al., 2011). Finally,our data result somewhat from similar to slightly higher than the range found at theNorth Hemisphere (red dashed lines; range: 1.5–1.7 ng m�3; Lindberg et al., 2007).The simple moving average of our data (SMA) is also reported (white line) in boththe graphs. (For interpretation of the references to colour in this figure legend, thereader is referred to the web version of this article.)
Fig. 5. Values for GEM found as a function of latitude over the Augusta basin. Alsoshown are compiled values for several marine/oceanic environmental systems (seetext for the references).
E. Bagnato et al. / Chemosphere 93 (2013) 2024–20
over the Yellow Sea (3.2–44 ng m�2 h�1; Ci et al.,Atlantic Ocean (20–80 ng m�2 h�1; Mason et al.,ontaining extremely high dissolved gaseous mer-aters. From Table 2it emerges that most of the
e Hg sea/air evasion fluxes have been calculatedng the gas-exchange model (GEM; Liss and Slater,
ly two data (plus the present study) refer to Hg fluxd by the dynamic flux chamber technique (DFC).dynamic flux chamber technique coupled with ac adsorption spectrometer (Lumex-RA 915+) thusportant step aimed to refine the estimation meth-fluxes from environmental surfaces (Wang et al.,nique also aims to reduce the uncertainty in thecessing data often given by the calculation model,ent from the choice of gas transfer parameteriza-of, 1992) and diffusion coefficient of mercury (Liss
). To evaluate the impact of Hg emissions from dif-e must have a clear understanding of the factorssions, develop a data base of emissions from sub-
ide range of Hg concentrations, and develop acaling point source measurements to broad areas.e 2, our data, and more in general Hg evasion ratesvironments, result to be higher than Hg flux from
ontaminated soils (�0.9 ng Hg m�2 h�1), while arehose reported from volcanic/geothermal areas andiated with hydrothermal systems (�14 and 83 ngectively; Table 2). The highest Hg fluxes measuredal activity are most likely due to a combination of
ing hydrothermal gas flow through soil and ele-ntrations in thermal area substrates (Varekamp4). Finally, Hg flux measured at the sea/air inter-sta Bay results to be several orders of magnitudeeleased from areas associated with important oretal mining, which are typically enriched in Hg rel-background concentrations (Table 2), and have
ranging from background rates (2 ng Hg m�2 h�1)ands of ng m�2 h�1 (3730–118000 ng Hg m�2 h�1;r to calculate the total sea–air Hg evasion flux overe area of the Augusta basin (about 23.5 km2), weof territorial distribution proposed by Aurenham-‘Voronoi Polygons’ method). This method allowedasin in seven different areas (ST1-7, in km2) eachdifferent % of the total Hg evasion flux. Thus, weulative Hg evasion flux for the whole basin of
yr�1 (�9.7 ± 0.1 g d�1), which accounts forglobal mercury oceanic evasion of 2000 t yr�1 pro-et al. (1994). Anyway, this value (9.7 ± 0.1 g d�1)
er than the total Hg flux emitted from the pollutedge 19–249 g d�1; Sakata et al., 2006; Narukawat it is significant if we consider that the extent ofe area of the Augusta basin represents only a trivialeanic surfaces on Earth (3.6 � 108 km2; Eakins andand 1/5 of the Tokyo Bay surface area (1000 km2).
ional flux assessment
ransport of atmospheric gaseous Hg, followed byposition, is an important process by which Hg isstrial and aquatic ecosystems far from its source.data collected during a very short-term survey
11 to April 2012) (range: 21–32 ng L�1; Appendixle to those reported for rainwaters collected at thecean (10–50 ng L�1; Nishimura, 1979) and theL�1; Cambray et al., 1979). We also found a good
Hg levels found in precipitations collected close to�1
32 2029
c electrolysis plant (industrial area; 17 ng L ) and
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Table 2Mercury evasion flux from some acquatic environments reported in literature including this study. For a more detailed description on averages and methods the reader is referredto the original article.
Measurement sites Date Hg(0) evasion flux (ng m�2 h�1) Methods References
(year-month-day) Range Mean (SD; n)
Augusta basinST1 29/11/2011 (35.6–36.3) 36 (0.3; 2963) DFC Present studyST2 29/11/2011 (14.2–14.5) 14.4 (0.1; 2965) DFC Present studyST3 30/11/2011 (71.8–72.1) 72 (0.1; 2958) DFC Present studyST4 29/11/2011 (3.2–3.9) 3.6 (0.3; 2963) DFC Present studyST5 24/06/2012 (10.4–11.1) 10.8 (0.3; 2752) DFC Present studyST6 23/06/2012 (7.1–7.3) 7.2 (0.1; 3244) DFC Present studyST7 25/06/2012 (17.8–18.2) 18 (0.2; 4293) DFC Present study
Other acquatic sitesEquatorial Pacific ocean 1984/07/03–1984/06/08 (0.5–8) 3 (2; 22) GEM Kim and Fitzgerald (1986)Western Mediterranean 2003/08/20–23 (4.1–6.2) 5.1 (1; 275) GEM Andersson et al. (2007)Western Mediterranean 2000/07/14–2000/08/09 (0.5–4.5) 2.5 (1.2; 6) GEM Gardfeldt et al. (2003)Eastern Mediterranean 2000/07/17–23 (1.6–15.2) 7.9 (4.2; 10) GEM Gardfeldt et al. (2003)Eastern Mediterranean 2010/08/26–2010/09/13 (0.2–4.9) 2.2 (1.5; 17) GEM Fantozzi et al. (2013)Mediterranean Sea 1998/02/06–1998/09/22 (1.2–5.7) 2.4 (1.5; 6) DFC Ferrara et al. (2000)Tyrrenian sea 2003/08/27–2004/10/29 (0.4–4.1) 1.6 (1.3; 675) GEM Andersson et al. (2007)Tyrrenian sea 2000/07/29–2000/08/08 (0.1–9.9) 4.2 (3.2; 7) GEM Gardfeldt et al. (2003)Ionian sea 2003/08/08–2004/11/11 (0.8–6.6) 2.7 (1.8; 888) GEM Andersson et al. (2007)Adriatic sea 2004/11/02–10 (2–9.7) 5.4 (2.5; 401) GEM Andersson et al. (2007)North Adriatic sea 2004/11/05–06 (23.7–33.2) 28.4 (4.7; 104) GEM Andersson et al. (2007)Strait of Sicily 2003/08/06–2004/03/26 (0.7–3.5) 2.1 (1.4; 329) GEM Andersson et al. (2007)Mediterranean coastal water 2000/07/31–2000/08/07 (2.7–4.5) 3.7 (0.8; 63) DFC Gardfeldt et al. (2003)North Atlantic Ocean 2005/07/07–11 (�0.6 to 2.5) 0.4 (0.3; 559) GEM Andersson et al. (2011)Baltic sea 1997/07/02–15 (6–89) 31 (25; 11) GEM Wängberg et al. (2001)Artic ocean 2005/07/13–2005/09/25 (n.a.) 2.4 (n.a.) GEM Andersson et al. (2008)North sea 1992/09/n.a. (2.4–46) 20 (13; 11) GEM Baeyens and Leermakers (1998)South China Sea 2007/08/11–27 (0.2–15.3) 4.5 (3.4; 40) GEM Fu et al. (2010)Tokyo Bay 2003/12/10–2005/01/12 (0.1–22) 5.8 (5; 22) GEM Narukawa et al. (2006)Yellow sea 2010/07/10–17 (3.2–44) 18.3 (11.8; 40) GEM Ci et al. (2011)
Land evasionBackground unpolluted soils (US) n.a. (0.3–0.8) 0.9 (0.2; 1326) DFC Ericksen et al. (2006)Volcanic/geothermal areas (LVC) 2000/04/14–15 (5.2–19.8) 13.7 (8; 12) DFC Engle and Gustin (2002)Mineralized area (Peavine peak, Nevada) 2000/04/14–15 (2–15) 10 (n.a.; 16) DFC Engle and Gustin (2002)Mine-waste enriched soils (Mt. Amiata) 2008/08/27–28 (250–8000) 3730 (n.a.; 56) DFC Fantozzi et al. (2013), in pressGold Mines (Venezuela) 2004/05/16–31 (650–420100) 118000 (n.a.; 12) DFC Garcıa-Sanchez et al. (2006)Hydrothermal systems (Lassen Park) 2004/08/20–21 (-110 to 103) 12 (n.a.; 13) DFC Engle et al. (2006)
n.a.;(n.a.)
D
2030 E. Bagnato et al. / Chemosphere 93 (2013) 2024–2032
the mineralized area of Mt. Amiata (Cinnabar deposits) near va-or-dominated geothermal springs (14.4 ng L�1) (Ferrara et al.,986). By attempting to calculate a first Hg bulk depositional flux
et + dry) for the Augusta basin, we used the following relation:
Hg ¼ ðCHg � PÞT�1 ð2Þ
here CHg is the concentration of Hg in rain (in ng L�1), P is themount of precipitation (in mm), and T is the exposition time ofe collector (in days). We estimated a preliminary Hg bulk depo-
itional flux ranging from 0.05 to 0.23 lg m�2 d�1 (weighted aver-ge of 0.10 lg m�2 d�1; Appendix II). Although our estimatedverage Hg bulk deposition flux (35.8 lg m�2 yr�1) at the Augustaasin is higher than the values calculated by Mason et al. (1994)
ocean (from 0.13 to 9.5 lg m�2 yr�1) and land (0.1–9.8 lg m�2 yr�1) at various latitudes (Downs et al., 1998), it re-ults to be one order of magnitude lower than the annually atmo-pheric Hg flux released in the MBL (maximum emission315 lg m�2 yr�1; this work).
. Conclusions
Mercury has an extremely complex cycle in the Earth’s ecosys-ms and the environmental bodies are both active sink and source
atmosphereand environposed in thstudy the pthe Augustaon Hg mass
Acknowled
This wothe Regionafully acknogusta andfor their lographic camto thank thographic cranalysis.
Appendix A
Supplemthe online
Hydrothermal systems (Yellowstone) 2003/09/12–2004/09/01 (-27 to 541) 83 (Sulfur Bank geothermal area n.a. (436–510) n.a
FC = dynamic flux chamber; GEM = gas-exchange model; n.a. = not available.
r Hg. The exchange of mercury between natural surfaces and the 2013.07.025.
an important process for the atmospheric cyclingtal turnover of this element. The new data set pro-
tudy offers a unique and original opportunity tontial outflow of Hg from the sea–air interface atsin, and will serve as a basis for future estimatesance in this area.
ents
s part of the IAMC-CNR/ASP project founded byeath Department of Syracuse. The authors grate-ge the personnel of the port authorities of Au-
F. Bulfamante from IAMC-CNR (Capo Granitola),ic support before, after and during the oceano-igns performed inside the Rada. They also wishchnical staff of IAMC-CNR involved in the ocean-s and Dr. F. Falco for her assistance during the
pplementary data
ry data associated with this article can be found, insion, at http://dx.doi.org/10.1016/j.chemosphere.
106) DFC Engle et al. (2006)DFC Gustin (2003)
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